The Marine Life Information Network

Global warming (extreme)

Extreme emission scenario (by the end of this century 2081-2100) benchmark of:

• A 5°C rise in SST and NBT (coastal to the shelf seas),

• A 6°C rise in surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf, and

• A 5°C rise in surface air temperature in intertidal habitats exclusive to Scotland. Further detail.

Evidence

Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas that frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek, 1992). In the north Atlantic, this species occurs from Norway to the coast of Spain. In the western Atlantic, Mytilus edulis has been observed to be expanding its range pole-wards and has reappeared in Svalbard, due to an increase in sea temperature in that region (Berge et al., 2005), whilst its equatorial limits have contracted approximately 350 km north of its previous southern limits in Cape Hatteras, North Carolina, due to increases in water temperature beyond the lethal limit (Jones et al., 2009).

Wells & Gray (1960) suggested that the mean summer water temperatures of 26.6 °C set the southern range limit. Gonzalez & Yevich (1976) found that Mytilus edulis could not tolerate sustained temperatures of 27°C, and feeding stopped after 25°C. Pearce (1969) found that whilst some populations of Mytilus edulis could survive at 27°C for over a month, none of these populations could survive temperatures of 28°C for more than four days. Read & Cummings (1967) estimated the upper tolerance limit to be 27°C, and Chapple et al. (1998) found that Mytilus edulis could not acclimate to temperatures above 28.5°C. Almada-Villela et al. (1982) found that growth significantly declined in juvenile Mytilus edulis as temperatures increased above 20°C. Similarly, Hiebenthal et al. (2013) found that the growth rate decreased by 60% as temperatures increased from 20 - 25°C and resulted in 25% mortality under experimental conditions. Incze et al. (1980) found that Mytilus edulis growth decreased at 20°C and mortality occurred at 25°C, although mortality occurred at lower temperatures when phytoplankton abundance was low, suggesting that mortality occurred through a combination of reducing food source at a time of metabolic stress. Lethal water temperatures appear to vary between areas (Tsuchiya, 1983) and it appears that tolerance varies, depending on the temperature range to which the individuals are acclimatised (Kittner & Riisgard, 2005). After acclimation of individuals of Mytilus edulis to 18°C, Kittner & Riisgaard (2005) observed that the filtration rates were at their maximum between 8.3 and 20°C and below this at 6°C the mussels closed their valves. However, after acclimation at 11°C for five days, the mussels maintained the high filtration rates down to 4°C.  Hence, given time, mussels can acclimatise, shifting their temperature tolerance.

Rising air temperatures can also lead to significant mortality in Mytilus edulis. Intertidal ecosystems are likely to be more negatively impacted than subtidal ecosystems, due to their increased daily and seasonal variations in temperatures (Jones et al., 2009). Tsuchiya (1983) documented the mass mortality of Mytilus edulis in August 1981 due to air temperatures of 34°C that resulted in mussel tissue temperatures over 40°C. In one hour, 50% of the Mytilus edulis from the upper 75% of the shore had died. It could not be concluded from this study whether the mortality was due to high temperatures, desiccation or a combination of the two. Under experimental conditions, exposure to air temperatures greater than 30°C led to significant mortality (Jones et al., 2009), suggesting this may be an upper temperature threshold for this species.

At the upper range of a mussel’s tolerance limit, heat shock proteins are produced, indicating high stress levels (Jones et al., 2010). After a single day at 30°C, heat shock proteins were still present over 14 days later, although at a reduced level. Increased temperatures can also affect reproduction in Mytilus edulis (Myrand et al., 2000). In shallow lagoons, mortality began in late July at the end of a major spawning event when temperatures peaked at >20°C. These mussels had a low energetic content post-spawning and had stopped shell growth.  The high temperatures likely caused mortality due to the reduced condition of the mussels post-spawning (Myrand et al., 2000). Gamete production does not appear to be affected by temperature (Suchanek, 1985).

Temperature changes may also lead to indirect effects. For example, an increase in temperature increases the mussels’ susceptibility to pathogens (Vibrio tubiashii) in the presence of relatively low concentrations of copper (Parry & Pipe, 2004).  Increased temperatures may also allow for range expansion of parasites or pathogens which will have a negative impact upon the health of the mussels if they become infected. There is evidence that increases in temperature will also give a competitive advantage to invasive species. For example, in the Dutch Wadden Sea mild winters favour Magallana gigas recruitment while cold winters favour Mytilus edulis (Deiderich, 2005).

Sensitivity assessment. Sea surface temperatures around the UK are currently between 6-19°C (Huthnance, 2010). Under the three scenarios (middle and high emission and extreme scenarios), summer sea temperatures in the south of the UK may rise to temperatures of 22, 23, and 24°C respectively. Mytilus edulis is a eurythermal species, and the maximum upper thermal limit of this species appears to generally be somewhere between 25 - 28°C, above which this species experiences mortality, with tolerance related to exposure. As ocean warming will occur gradually, across the course of this century, it is expected that this species will be able to withstand these increases in temperature.

As this species is intertidal, it will also have to cope with increases air temperature. In July, temperatures can reach up to an average of 25°C in the south of the UK, although the highest temperature recorded 1961-2010 was 38.5°C (Perry & Golding, 2011). If air temperatures rise by 3, 4, and 6°C by the end of the century (middle and high emission and extreme scenarios, respectively), this could lead to temperatures reaching average summer high temperatures of between 28 - 31°C.

Under the middle and high emission scenario with seawater temperatures reaching up to 23°C and air temperatures reaching 29°C, Mytilus edulis may be able to adapt to global warming.  Most studies place its upper thermal limit at between 25-28°C for seawater temperatures and 30°C for air temperature, although these temperatures may lead to a decrease in growth. Under these scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. Therefore, this biotope is assessed as ‘Not sensitive’ to ocean warming under the middle and high emission scenarios.

Under the extreme scenario, whilst seawater temperatures are expected to remain below the threshold upper temperature limit for this species, air temperatures are likely to rise to 31°C, which exceeds potential upper air temperature limits, and is likely to lead to some mortality in the south of the UK. As such, resistance has been assessed as ‘Medium’, whilst resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ to ocean warming under the extreme scenario.

Mytilus galloprovincialis is a closely related species to Mytilus edulis, which occurs around the coastline of SW England, SW Ireland, and South Wales. Mytilus galloprovincialis has a higher thermal limit than Mytilus edulis; up to 31°C dependent on acclimation temperature (Braby & Somero, 2006), so it is potentially feasible that Mytilus galloprovincialis may replace blue mussels in the UK with gradual global warming, although there is no evidence to support this hypothesis at present. Furthermore, Mytilus edulis hybridizes with Mytilus galloprovincialis where their ranges overlap in the UK, although this does not appear to confer any temperature advantage, as the physiology of Mytilus edulis and the hybrids are very similar (Beaumont et al., 2004).

Global warming (high)

High emission scenario (by the end of this century 2081-2100) benchmark of:

• A 4°C rise in SST, NBT (coastal to the shelf seas) and surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf, and

• A 3°C rise in surface air temperature in intertidal habitats exclusive to Scotland. Further detail.

Evidence

Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas that frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek, 1992). In the north Atlantic, this species occurs from Norway to the coast of Spain. In the western Atlantic, Mytilus edulis has been observed to be expanding its range pole-wards and has reappeared in Svalbard, due to an increase in sea temperature in that region (Berge et al., 2005), whilst its equatorial limits have contracted approximately 350 km north of its previous southern limits in Cape Hatteras, North Carolina, due to increases in water temperature beyond the lethal limit (Jones et al., 2009).

Wells & Gray (1960) suggested that the mean summer water temperatures of 26.6 °C set the southern range limit. Gonzalez & Yevich (1976) found that Mytilus edulis could not tolerate sustained temperatures of 27°C, and feeding stopped after 25°C. Pearce (1969) found that whilst some populations of Mytilus edulis could survive at 27°C for over a month, none of these populations could survive temperatures of 28°C for more than four days. Read & Cummings (1967) estimated the upper tolerance limit to be 27°C, and Chapple et al. (1998) found that Mytilus edulis could not acclimate to temperatures above 28.5°C. Almada-Villela et al. (1982) found that growth significantly declined in juvenile Mytilus edulis as temperatures increased above 20°C. Similarly, Hiebenthal et al. (2013) found that the growth rate decreased by 60% as temperatures increased from 20 - 25°C and resulted in 25% mortality under experimental conditions. Incze et al. (1980) found that Mytilus edulis growth decreased at 20°C and mortality occurred at 25°C, although mortality occurred at lower temperatures when phytoplankton abundance was low, suggesting that mortality occurred through a combination of reducing food source at a time of metabolic stress. Lethal water temperatures appear to vary between areas (Tsuchiya, 1983) and it appears that tolerance varies, depending on the temperature range to which the individuals are acclimatised (Kittner & Riisgard, 2005). After acclimation of individuals of Mytilus edulis to 18°C, Kittner & Riisgaard (2005) observed that the filtration rates were at their maximum between 8.3 and 20°C and below this at 6°C the mussels closed their valves. However, after acclimation at 11°C for five days, the mussels maintained the high filtration rates down to 4°C.  Hence, given time, mussels can acclimatise, shifting their temperature tolerance.

Rising air temperatures can also lead to significant mortality in Mytilus edulis. Intertidal ecosystems are likely to be more negatively impacted than subtidal ecosystems, due to their increased daily and seasonal variations in temperatures (Jones et al., 2009). Tsuchiya (1983) documented the mass mortality of Mytilus edulis in August 1981 due to air temperatures of 34°C that resulted in mussel tissue temperatures over 40°C. In one hour, 50% of the Mytilus edulis from the upper 75% of the shore had died. It could not be concluded from this study whether the mortality was due to high temperatures, desiccation or a combination of the two. Under experimental conditions, exposure to air temperatures greater than 30°C led to significant mortality (Jones et al., 2009), suggesting this may be an upper temperature threshold for this species.

At the upper range of a mussel’s tolerance limit, heat shock proteins are produced, indicating high stress levels (Jones et al., 2010). After a single day at 30°C, heat shock proteins were still present over 14 days later, although at a reduced level. Increased temperatures can also affect reproduction in Mytilus edulis (Myrand et al., 2000). In shallow lagoons, mortality began in late July at the end of a major spawning event when temperatures peaked at >20°C. These mussels had a low energetic content post-spawning and had stopped shell growth.  The high temperatures likely caused mortality due to the reduced condition of the mussels post-spawning (Myrand et al., 2000). Gamete production does not appear to be affected by temperature (Suchanek, 1985).

Temperature changes may also lead to indirect effects. For example, an increase in temperature increases the mussels’ susceptibility to pathogens (Vibrio tubiashii) in the presence of relatively low concentrations of copper (Parry & Pipe, 2004).  Increased temperatures may also allow for range expansion of parasites or pathogens which will have a negative impact upon the health of the mussels if they become infected. There is evidence that increases in temperature will also give a competitive advantage to invasive species. For example, in the Dutch Wadden Sea mild winters favour Magallana gigas recruitment while cold winters favour Mytilus edulis (Deiderich, 2005).

Sensitivity assessment. Sea surface temperatures around the UK are currently between 6-19°C (Huthnance, 2010). Under the three scenarios (middle and high emission and extreme scenarios), summer sea temperatures in the south of the UK may rise to temperatures of 22, 23, and 24°C respectively. Mytilus edulis is a eurythermal species, and the maximum upper thermal limit of this species appears to generally be somewhere between 25 - 28°C, above which this species experiences mortality, with tolerance related to exposure. As ocean warming will occur gradually, across the course of this century, it is expected that this species will be able to withstand these increases in temperature.

As this species is intertidal, it will also have to cope with increases air temperature. In July, temperatures can reach up to an average of 25°C in the south of the UK, although the highest temperature recorded 1961-2010 was 38.5°C (Perry & Golding, 2011). If air temperatures rise by 3, 4, and 6°C by the end of the century (middle and high emission and extreme scenarios, respectively), this could lead to temperatures reaching average summer high temperatures of between 28 - 31°C.

Under the middle and high emission scenario with seawater temperatures reaching up to 23°C and air temperatures reaching 29°C, Mytilus edulis may be able to adapt to global warming.  Most studies place its upper thermal limit at between 25-28°C for seawater temperatures and 30°C for air temperature, although these temperatures may lead to a decrease in growth. Under these scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. Therefore, this biotope is assessed as ‘Not sensitive’ to ocean warming under the middle and high emission scenarios.

Under the extreme scenario, whilst seawater temperatures are expected to remain below the threshold upper temperature limit for this species, air temperatures are likely to rise to 31°C, which exceeds potential upper air temperature limits, and is likely to lead to some mortality in the south of the UK. As such, resistance has been assessed as ‘Medium’, whilst resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ to ocean warming under the extreme scenario.

Mytilus galloprovincialis is a closely related species to Mytilus edulis, which occurs around the coastline of SW England, SW Ireland, and South Wales. Mytilus galloprovincialis has a higher thermal limit than Mytilus edulis; up to 31°C dependent on acclimation temperature (Braby & Somero, 2006), so it is potentially feasible that Mytilus galloprovincialis may replace blue mussels in the UK with gradual global warming, although there is no evidence to support this hypothesis at present. Furthermore, Mytilus edulis hybridizes with Mytilus galloprovincialis where their ranges overlap in the UK, although this does not appear to confer any temperature advantage, as the physiology of Mytilus edulis and the hybrids are very similar (Beaumont et al., 2004).

Global warming (middle)

Middle emission scenario (by the end of this century 2081-2100) benchmark of:

• A 3°C rise in SST, NBT (coastal to the shelf seas) and surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf.

• A 2°C rise in surface air temperature in intertidal habitats exclusive to Scotland. Further detail.

Evidence

Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas that frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek, 1992). In the north Atlantic, this species occurs from Norway to the coast of Spain. In the western Atlantic, Mytilus edulis has been observed to be expanding its range pole-wards and has reappeared in Svalbard, due to an increase in sea temperature in that region (Berge et al., 2005), whilst its equatorial limits have contracted approximately 350 km north of its previous southern limits in Cape Hatteras, North Carolina, due to increases in water temperature beyond the lethal limit (Jones et al., 2009).

Wells & Gray (1960) suggested that the mean summer water temperatures of 26.6 °C set the southern range limit. Gonzalez & Yevich (1976) found that Mytilus edulis could not tolerate sustained temperatures of 27°C, and feeding stopped after 25°C. Pearce (1969) found that whilst some populations of Mytilus edulis could survive at 27°C for over a month, none of these populations could survive temperatures of 28°C for more than four days. Read & Cummings (1967) estimated the upper tolerance limit to be 27°C, and Chapple et al. (1998) found that Mytilus edulis could not acclimate to temperatures above 28.5°C. Almada-Villela et al. (1982) found that growth significantly declined in juvenile Mytilus edulis as temperatures increased above 20°C. Similarly, Hiebenthal et al. (2013) found that the growth rate decreased by 60% as temperatures increased from 20 - 25°C and resulted in 25% mortality under experimental conditions. Incze et al. (1980) found that Mytilus edulis growth decreased at 20°C and mortality occurred at 25°C, although mortality occurred at lower temperatures when phytoplankton abundance was low, suggesting that mortality occurred through a combination of reducing food source at a time of metabolic stress. Lethal water temperatures appear to vary between areas (Tsuchiya, 1983) and it appears that tolerance varies, depending on the temperature range to which the individuals are acclimatised (Kittner & Riisgard, 2005). After acclimation of individuals of Mytilus edulis to 18°C, Kittner & Riisgaard (2005) observed that the filtration rates were at their maximum between 8.3 and 20°C and below this at 6°C the mussels closed their valves. However, after acclimation at 11°C for five days, the mussels maintained the high filtration rates down to 4°C.  Hence, given time, mussels can acclimatise, shifting their temperature tolerance.

Rising air temperatures can also lead to significant mortality in Mytilus edulis. Intertidal ecosystems are likely to be more negatively impacted than subtidal ecosystems, due to their increased daily and seasonal variations in temperatures (Jones et al., 2009). Tsuchiya (1983) documented the mass mortality of Mytilus edulis in August 1981 due to air temperatures of 34°C that resulted in mussel tissue temperatures over 40°C. In one hour, 50% of the Mytilus edulis from the upper 75% of the shore had died. It could not be concluded from this study whether the mortality was due to high temperatures, desiccation or a combination of the two. Under experimental conditions, exposure to air temperatures greater than 30°C led to significant mortality (Jones et al., 2009), suggesting this may be an upper temperature threshold for this species.

At the upper range of a mussel’s tolerance limit, heat shock proteins are produced, indicating high stress levels (Jones et al., 2010). After a single day at 30°C, heat shock proteins were still present over 14 days later, although at a reduced level. Increased temperatures can also affect reproduction in Mytilus edulis (Myrand et al., 2000). In shallow lagoons, mortality began in late July at the end of a major spawning event when temperatures peaked at >20°C. These mussels had a low energetic content post-spawning and had stopped shell growth.  The high temperatures likely caused mortality due to the reduced condition of the mussels post-spawning (Myrand et al., 2000). Gamete production does not appear to be affected by temperature (Suchanek, 1985).

Temperature changes may also lead to indirect effects. For example, an increase in temperature increases the mussels’ susceptibility to pathogens (Vibrio tubiashii) in the presence of relatively low concentrations of copper (Parry & Pipe, 2004).  Increased temperatures may also allow for range expansion of parasites or pathogens which will have a negative impact upon the health of the mussels if they become infected. There is evidence that increases in temperature will also give a competitive advantage to invasive species. For example, in the Dutch Wadden Sea mild winters favour Magallana gigas recruitment while cold winters favour Mytilus edulis (Deiderich, 2005).

Sensitivity assessment. Sea surface temperatures around the UK are currently between 6-19°C (Huthnance, 2010). Under the three scenarios (middle and high emission and extreme scenarios), summer sea temperatures in the south of the UK may rise to temperatures of 22, 23, and 24°C respectively. Mytilus edulis is a eurythermal species, and the maximum upper thermal limit of this species appears to generally be somewhere between 25 - 28°C, above which this species experiences mortality, with tolerance related to exposure. As ocean warming will occur gradually, across the course of this century, it is expected that this species will be able to withstand these increases in temperature.

As this species is intertidal, it will also have to cope with increases air temperature. In July, temperatures can reach up to an average of 25°C in the south of the UK, although the highest temperature recorded 1961-2010 was 38.5°C (Perry & Golding, 2011). If air temperatures rise by 3, 4, and 6°C by the end of the century (middle and high emission and extreme scenarios, respectively), this could lead to temperatures reaching average summer high temperatures of between 28 - 31°C.

Under the middle and high emission scenario with seawater temperatures reaching up to 23°C and air temperatures reaching 29°C, Mytilus edulis may be able to adapt to global warming.  Most studies place its upper thermal limit at between 25-28°C for seawater temperatures and 30°C for air temperature, although these temperatures may lead to a decrease in growth. Under these scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. Therefore, this biotope is assessed as ‘Not sensitive’ to ocean warming under the middle and high emission scenarios.

Under the extreme scenario, whilst seawater temperatures are expected to remain below the threshold upper temperature limit for this species, air temperatures are likely to rise to 31°C, which exceeds potential upper air temperature limits, and is likely to lead to some mortality in the south of the UK. As such, resistance has been assessed as ‘Medium’, whilst resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ to ocean warming under the extreme scenario.

Mytilus galloprovincialis is a closely related species to Mytilus edulis, which occurs around the coastline of SW England, SW Ireland, and South Wales. Mytilus galloprovincialis has a higher thermal limit than Mytilus edulis; up to 31°C dependent on acclimation temperature (Braby & Somero, 2006), so it is potentially feasible that Mytilus galloprovincialis may replace blue mussels in the UK with gradual global warming, although there is no evidence to support this hypothesis at present. Furthermore, Mytilus edulis hybridizes with Mytilus galloprovincialis where their ranges overlap in the UK, although this does not appear to confer any temperature advantage, as the physiology of Mytilus edulis and the hybrids are very similar (Beaumont et al., 2004).

Marine heatwaves (high)

High emission scenario benchmark: A marine heatwave occurring every two years, with a mean duration of 120 days, and a maximum intensity of 3.5°C. Further detail.

Evidence

Marine heatwaves due to increased air-sea heat flux are predicted to occur more frequently, at an increased intensity and last for longer by the end of this century under both middle and high emission scenarios (Frölicher et al., 2018). Intertidal populations of Mytilus edulis may be particularly sensitive to marine heatwaves. When submerged, a mussel’s body temperature closely approximates that of the surrounding water, whereas when emerged, body temperatures can become much higher than the surrounding air or substratum (Zippay & Helmuth, 2012). In the southern portion of its range in the USA, intertidal populations of Mytilus edulis have experienced catastrophic mortality directly associated with summer high temperatures of up to 32°C, with populations shifting their range 350 km northwards of their previous range (Jones et al., 2009).

Air temperatures tend to be more variable and extreme than seawater temperatures (Helmuth et al., 2002). While the south of the UK has a mean summer daily high temperature of 21°C, temperatures can often reach ≥30°C (Met Office, 2016). Temperature loggers on the west coast of Scotland recorded intertidal temperatures on the high shore exceeding 40°C in seven of the 11 years it was monitored (Burrows, 2017), which shows the extreme temperatures that intertidal species have to cope with, at present. Furthermore, when exposed to high daytime temperatures, internal body temperature can far exceed air temperatures. For example, when Mytilus edulis was exposed to air temperatures of up to 34°C on the shore, body temperatures of the mussels increased to 46°C, leading to mortality (Tsuchiya, 1983).

The thermal tolerance of Mytilus edulis decreases under repeated heat stress, therefore this species is likely to be especially sensitive to both marine and aerial heatwaves (Seuront et al., 2019). In Japan, in 1981, mass mortality of Mytilus edulis occurred along a rocky shore as a result of unusually high temperatures, whilst another species of mussel (Mytilisepta virgatus) which occurred in the zone above Mytilus edulis exhibited much greater levels of heat tolerance and low mortality (Tsuchiya, 1983). Mytilus edulis is thought to be particularly susceptible to high temperatures and heatwaves in the summer, due to the low energy reserves of the organism after spawning (Tremblay et al., 1998, Myrand et al., 2000).  

Sensitivity Assessment. Under the middle emission scenario, if heatwaves occurred every three years, with a maximum intensity of 2°C for 80 days by the end of this century, this could lead to summer sea temperatures reaching up to 24°C in southern England. Mytilus edulis is thought to have an upper thermal limit of 25-28°C, although growth decreases at temperatures over 20°C (see Global Warming).  As this biotope occurs in the intertidal, this species will not only experience increased sea surface temperatures but will also experience extreme air temperatures. Whilst this species may be able to cope with sea temperatures reaching 24°C, it is likely that during exposure it will be subjected to temperatures exceeding 30°C, which will lead to some mortality, particularly in the south. Therefore, resistance has been assessed as ‘Medium’. As a further heatwave is likely to affect this habitat before full recovery, recovery has been assessed as ‘Very Low.’ Therefore, this biotope is assessed as having ‘Medium’ sensitivity to marine heatwaves under the middle emission scenario.

Under the high emission scenario, if heatwaves occur every two years by the end of this century, reaching a maximum intensity of 3.5°C for 120 days, this could lead to the heatwave lasting the entire summer with temperatures reaching up to 26.5°C, and air temperatures exceeding 30°C across much of the UK. This is likely to cause significant mortality and, therefore, resistance has been assessed as ‘Low’. As a further heatwave is likely to affect this habitat before full recovery, recovery has been assessed as ‘Very Low.’ Therefore, this biotope is assessed as having ‘High’ sensitivity to marine heatwaves under the high emission scenario.

Marine heatwaves (middle)

Middle emission scenario benchmark:  A marine heatwave occurring every three years, with a mean duration of 80 days, with a maximum intensity of 2°C. Further detail.

Evidence

Marine heatwaves due to increased air-sea heat flux are predicted to occur more frequently, at an increased intensity and last for longer by the end of this century under both middle and high emission scenarios (Frölicher et al., 2018). Intertidal populations of Mytilus edulis may be particularly sensitive to marine heatwaves. When submerged, a mussel’s body temperature closely approximates that of the surrounding water, whereas when emerged, body temperatures can become much higher than the surrounding air or substratum (Zippay & Helmuth, 2012). In the southern portion of its range in the USA, intertidal populations of Mytilus edulis have experienced catastrophic mortality directly associated with summer high temperatures of up to 32°C, with populations shifting their range 350 km northwards of their previous range (Jones et al., 2009).

Air temperatures tend to be more variable and extreme than seawater temperatures (Helmuth et al., 2002). While the south of the UK has a mean summer daily high temperature of 21°C, temperatures can often reach ≥30°C (Met Office, 2016). Temperature loggers on the west coast of Scotland recorded intertidal temperatures on the high shore exceeding 40°C in seven of the 11 years it was monitored (Burrows, 2017), which shows the extreme temperatures that intertidal species have to cope with, at present. Furthermore, when exposed to high daytime temperatures, internal body temperature can far exceed air temperatures. For example, when Mytilus edulis was exposed to air temperatures of up to 34°C on the shore, body temperatures of the mussels increased to 46°C, leading to mortality (Tsuchiya, 1983).

The thermal tolerance of Mytilus edulis decreases under repeated heat stress, therefore this species is likely to be especially sensitive to both marine and aerial heatwaves (Seuront et al., 2019). In Japan, in 1981, mass mortality of Mytilus edulis occurred along a rocky shore as a result of unusually high temperatures, whilst another species of mussel (Mytilisepta virgatus) which occurred in the zone above Mytilus edulis exhibited much greater levels of heat tolerance and low mortality (Tsuchiya, 1983). Mytilus edulis is thought to be particularly susceptible to high temperatures and heatwaves in the summer, due to the low energy reserves of the organism after spawning (Tremblay et al., 1998, Myrand et al., 2000).  

Sensitivity Assessment. Under the middle emission scenario, if heatwaves occurred every three years, with a maximum intensity of 2°C for 80 days by the end of this century, this could lead to summer sea temperatures reaching up to 24°C in southern England. Mytilus edulis is thought to have an upper thermal limit of 25-28°C, although growth decreases at temperatures over 20°C (see Global Warming).  As this biotope occurs in the intertidal, this species will not only experience increased sea surface temperatures but will also experience extreme air temperatures. Whilst this species may be able to cope with sea temperatures reaching 24°C, it is likely that during exposure it will be subjected to temperatures exceeding 30°C, which will lead to some mortality, particularly in the south. Therefore, resistance has been assessed as ‘Medium’. As a further heatwave is likely to affect this habitat before full recovery, recovery has been assessed as ‘Very Low.’ Therefore, this biotope is assessed as having ‘Medium’ sensitivity to marine heatwaves under the middle emission scenario.

Under the high emission scenario, if heatwaves occur every two years by the end of this century, reaching a maximum intensity of 3.5°C for 120 days, this could lead to the heatwave lasting the entire summer with temperatures reaching up to 26.5°C, and air temperatures exceeding 30°C across much of the UK. This is likely to cause significant mortality and, therefore, resistance has been assessed as ‘Low’. As a further heatwave is likely to affect this habitat before full recovery, recovery has been assessed as ‘Very Low.’ Therefore, this biotope is assessed as having ‘High’ sensitivity to marine heatwaves under the high emission scenario.

Ocean acidification (high)

High emission scenario benchmark: a further decrease in pH of 0.35 (annual mean) and corresponding 120% increase in H+ ions , seasonal aragonite saturation of 20% of UK coastal waters and North Sea bottom waters, and the aragonite saturation horizon in the NE Atlantic, off the continental shelf, occurring at a depth of 400 m by the end of this century 2081-2100. Further detail 

Evidence

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005), with it expected to drop up to a further 0.35 units by the end of this century, depending on emission scenario. In general, it is thought that calcifying invertebrates will be more sensitive to ocean acidification than non-calcifying invertebrates, which appear to have a more mixed response (Hofmann et al., 2010). It must be noted that many species show variation in their response to pCO₂ independent of their taxonomic group or habitat preferences (Widdicombe & Spicer, 2008; Kroeker et al., 2013).

Mytilus edulis is a calcified organism but it is unlikely that this species will be significantly negatively impacted by ocean acidification, because acidification does not appear to lead to mortality, even at levels which far exceed levels of ocean acidification expected for the end of this century (e.g. Berge et al., 2006, Melzner et al., 2011). For example, levels of growth in Mytilus edulis were maintained at pH 7.6 -7.7, although growth does decrease under pH levels < 7.4 (Berge et al., 2006, Melzner et al., 2011).

The calcified shell of Mytilus edulis consists of an outer calcite layer and an inner aragonite layer (Fitzer et al., 2015). When cultured at levels of acidification expected for the end of this century, under both the middle (550 ppm) and high (1000 ppm) emission scenario, results showed that Mytilus edulis shells became more brittle (Fitzer et al., 2015). There was no impact of ocean acidification on the production or strength of the byssal threads (Dickey et al., 2018). Beesley et al. (2008) found that the health of Mytilus edulis decreased as a result of 60 days of exposure to increased CO₂, which they suggested was due to the elevated concentration of calcium ions in the haemolymph.  Sun et al. (2017) found that ocean acidification damaged the ultrastructure of haemocytes and led to a reduction in phagocytosis.

The Baltic Sea is naturally low in carbonates, and exhibits seasonal aragonite undersaturation and borderline calcite undersaturation, even at levels of low pCO₂ (Thomsen & Melzner, 2010), yet Mytilus edulis is one of the most conspicuous animals present on the rocky sublittoral of the northern Baltic (Westerbom et al., 2002). In Kiels Fjord in the Baltic Sea, where pH can reach levels of <7.5 in the summer, Mytilus edulis can be found in high densities, with juvenile settlement occurring in the summer when pH values are at their lowest (Thomsen et al., 2010). This suggests that this species will be able to tolerate pH levels expected for the end of this century around the UK.

Sensitivity Assessment. Whilst levels of ocean acidification expected for the end of this century appear to decrease organism health through alterations to immune response, and an increase in shell brittleness, it is not certain how these impacts will lead to population-level responses. In situ data show that the piddock Barnea candida and the mussel Mytilus edulis are present in the Baltic Sea, where pH levels can fluctuate from <7.5 - >8.2  (Thomsen et al., 2010) highlighting their ability to tolerate a wide pH range. Therefore, under both the middle and high emission scenario resistance is assessed as ‘High’, and resilience is assessed as ‘High’ leading to a score of ‘Not sensitive’ at the pressure benchmark.

Ocean acidification (middle)

Middle emission scenario benchmark: a further decrease in pH of 0.15 (annual mean) and corresponding 35% increase in H+ ions with no coastal aragonite undersaturation and the aragonite saturation horizon in the NE Atlantic, off the continental shelf, at a depth of 800 m by the end of this century 2081-2100. Further detail.

Evidence

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005), with it expected to drop up to a further 0.35 units by the end of this century, depending on emission scenario. In general, it is thought that calcifying invertebrates will be more sensitive to ocean acidification than non-calcifying invertebrates, which appear to have a more mixed response (Hofmann et al., 2010). It must be noted that many species show variation in their response to pCO₂ independent of their taxonomic group or habitat preferences (Widdicombe & Spicer, 2008; Kroeker et al., 2013).

Mytilus edulis is a calcified organism but it is unlikely that this species will be significantly negatively impacted by ocean acidification, because acidification does not appear to lead to mortality, even at levels which far exceed levels of ocean acidification expected for the end of this century (e.g. Berge et al., 2006, Melzner et al., 2011). For example, levels of growth in Mytilus edulis were maintained at pH 7.6 -7.7, although growth does decrease under pH levels < 7.4 (Berge et al., 2006, Melzner et al., 2011).

The calcified shell of Mytilus edulis consists of an outer calcite layer and an inner aragonite layer (Fitzer et al., 2015). When cultured at levels of acidification expected for the end of this century, under both the middle (550 ppm) and high (1000 ppm) emission scenario, results showed that Mytilus edulis shells became more brittle (Fitzer et al., 2015). There was no impact of ocean acidification on the production or strength of the byssal threads (Dickey et al., 2018). Beesley et al. (2008) found that the health of Mytilus edulis decreased as a result of 60 days of exposure to increased CO₂, which they suggested was due to the elevated concentration of calcium ions in the haemolymph.  Sun et al. (2017) found that ocean acidification damaged the ultrastructure of haemocytes and led to a reduction in phagocytosis.

The Baltic Sea is naturally low in carbonates, and exhibits seasonal aragonite undersaturation and borderline calcite undersaturation, even at levels of low pCO₂ (Thomsen & Melzner, 2010), yet Mytilus edulis is one of the most conspicuous animals present on the rocky sublittoral of the northern Baltic (Westerbom et al., 2002). In Kiels Fjord in the Baltic Sea, where pH can reach levels of <7.5 in the summer, Mytilus edulis can be found in high densities, with juvenile settlement occurring in the summer when pH values are at their lowest (Thomsen et al., 2010). This suggests that this species will be able to tolerate pH levels expected for the end of this century around the UK.

Sensitivity Assessment. Whilst levels of ocean acidification expected for the end of this century appear to decrease organism health through alterations to immune response, and an increase in shell brittleness, it is not certain how these impacts will lead to population-level responses. In situ data show that the piddock Barnea candida and the mussel Mytilus edulis are present in the Baltic Sea, where pH levels can fluctuate from <7.5 - >8.2  (Thomsen et al., 2010) highlighting their ability to tolerate a wide pH range. Therefore, under both the middle and high emission scenario resistance is assessed as ‘High’, and resilience is assessed as ‘High’ leading to a score of ‘Not sensitive’ at the pressure benchmark.

Sea level rise (extreme)

Extreme scenario benchmark: a 107 cm rise in average UK by the end of this century (2018-2100). Further detail.

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 1996, Crooks, 2004, Fujii & Raffaelli, 2008).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position with the coastal energy gradient (Crooks, 2004, Fujii & Raffaelli, 2008).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (1996) suggested that a sea-level rise rate of 6 mm/yr could result in landward movement of estuaries by 10 m/yr, long-shore migration of open coast landforms of 50 m/yr and ebb-tidal deltas to extend laterally by 300 m/yr. 

The effects of sea-level rise and increased wave action may be increased further due to storms and storm surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Lowe et al., 2018, Palmer et al., 2018).

This biotope (LS.LBR.LMus.Myt) occurs on the lower and mid-shore of the intertidal zone and, therefore, an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. Beds may be able to expand their range and migrate upwards to compensate for sea-level rise, if not constrained by lack of suitable habitat (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Mytilus edulis beds on littoral sediment will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. In this assessment, as the ability to migrate inshore will be site-specific, we will assess on a worst-case-scenario basis, assuming that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), more than half of this biotope may be lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small proportion of this biotope may be lost. Under the middle emission scenario, resistance has been assessed as ‘Medium’, as less than 25% of this biotope may be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘Medium’.

Under the high and extreme emission scenario, resistance has been assessed as ‘Low’, as more than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘High’.

Sea level rise (high)

High emission scenario benchmark: a 70 cm rise in average UK by the end of this century (2018-2100). Further detail.

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 1996, Crooks, 2004, Fujii & Raffaelli, 2008).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position with the coastal energy gradient (Crooks, 2004, Fujii & Raffaelli, 2008).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (1996) suggested that a sea-level rise rate of 6 mm/yr could result in landward movement of estuaries by 10 m/yr, long-shore migration of open coast landforms of 50 m/yr and ebb-tidal deltas to extend laterally by 300 m/yr. 

The effects of sea-level rise and increased wave action may be increased further due to storms and storm surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Lowe et al., 2018, Palmer et al., 2018).

This biotope (LS.LBR.LMus.Myt) occurs on the lower and mid-shore of the intertidal zone and, therefore, an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. Beds may be able to expand their range and migrate upwards to compensate for sea-level rise, if not constrained by lack of suitable habitat (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Mytilus edulis beds on littoral sediment will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. In this assessment, as the ability to migrate inshore will be site-specific, we will assess on a worst-case-scenario basis, assuming that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), more than half of this biotope may be lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small proportion of this biotope may be lost. Under the middle emission scenario, resistance has been assessed as ‘Medium’, as less than 25% of this biotope may be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘Medium’.

Under the high and extreme emission scenario, resistance has been assessed as ‘Low’, as more than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘High’.

Sea level rise (middle)

Middle emission scenario benchmark: a 50 cm rise in average UK sea-level rise by the end of this century (2081-2100). Further detail.

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 1996, Crooks, 2004, Fujii & Raffaelli, 2008).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position with the coastal energy gradient (Crooks, 2004, Fujii & Raffaelli, 2008).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (1996) suggested that a sea-level rise rate of 6 mm/yr could result in landward movement of estuaries by 10 m/yr, long-shore migration of open coast landforms of 50 m/yr and ebb-tidal deltas to extend laterally by 300 m/yr. 

The effects of sea-level rise and increased wave action may be increased further due to storms and storm surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Lowe et al., 2018, Palmer et al., 2018).

This biotope (LS.LBR.LMus.Myt) occurs on the lower and mid-shore of the intertidal zone and, therefore, an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. Beds may be able to expand their range and migrate upwards to compensate for sea-level rise, if not constrained by lack of suitable habitat (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Mytilus edulis beds on littoral sediment will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. In this assessment, as the ability to migrate inshore will be site-specific, we will assess on a worst-case-scenario basis, assuming that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), more than half of this biotope may be lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small proportion of this biotope may be lost. Under the middle emission scenario, resistance has been assessed as ‘Medium’, as less than 25% of this biotope may be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘Medium’.

Under the high and extreme emission scenario, resistance has been assessed as ‘Low’, as more than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘High’.

Temperature increase (local)

Benchmark. A 5°C increase in temperature for one month, or 2°C for one year. Further detail

Evidence

Local populations may be acclimated to the prevailing temperature regime and may, therefore, exhibit different tolerances to other populations subject to different salinity conditions and therefore caution should be used when inferring tolerances from populations in different regions. Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas which frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek, 1992).  In recent years, Mytilus edulis has been observed to be expanding its range pole-wards and has reappeared in Svalbard, due to an increase in sea temperature in that region (Berge et al., 2005), whilst its equatorial limits are contracting due to increases in water temperature beyond the lethal limit (Jones et al., 2010).  In British waters, 29°C was recorded as the upper sustained thermal tolerance limit for Mytilus edulis (Read & Cumming, 1967; Almada-Villela, et al., 1982), although it is thought that European mussels will rarely experience temperatures above 25°C (Seed & Suchanek, 1992). 

Tsuchiya (1983) documented the mass mortality of Mytilus edulis in in Mutsu Bay, northern Japan in August 1981 due to air temperatures of 34°C that resulted in mussel tissue temperatures in excess of 40°C.  In one hour, 50% of the Mytilus edulis from the upper 75% of the shore had died.  It could not be concluded from this study whether the mortality was due to high temperatures, desiccation or a combination of the two.  Lethal water temperatures appear to vary between areas (Tsuchiya, 1983) although it appears that their tolerance at certain temperatures vary, depending on the temperature range to which the individuals are acclimatised (Kittner & Riisgaard, 2005).  After acclimation of individuals of M. edulis to 18°C, Kittner & Riisgaard (2005) observed that the filtrations rates were at their maximum between 8.3 and 20°C and below this at 6°C the mussels closed their valves.  However, after being acclimated at 11°C for five days, the mussels maintained the high filtration rates down to 4°C.  Hence, given time, mussels can acclimatise and shift their temperature tolerance.  Filtration in Mytilus edulis was observed to continue down to -1°C, with high absorption efficiencies (53-81%) (Loo, 1992).

At the upper range of a mussels tolerance limit, heat shock proteins are produced, indicating high stress levels (Jones et al., 2010).  After a single day at 30°C, the heat shock proteins were still present over 14 days later, although at a reduced level.  Increased temperatures can affect reproduction in Mytilus edulis (Myrand et al., 2000).  In shallow lagoons, mortality began in late July at the end of a major spawning event when temperatures peaked at >20°C.  These mussels had a low energetic content post-spawning and had stopped shell growth.  It is likely that the high temperatures caused mortality due to the reduced condition of the mussels post-spawning (Myrand et al., 2000). Gamete production does not appear to be affected by temperature (Suchanek, 1985).

Temperature changes may also lead to indirect effects.  For example, an increase in temperature increases the mussels’ susceptibility to pathogens (Vibrio tubiashii) in the presence of relatively low concentrations of copper (Parry & Pipe, 2004).  Increased temperatures may also allow for range expansion of parasites or pathogens which will have a negative impact on the health of the mussels if they become infected.

Power stations have the potential to cause an increase in sea temperature of up to 15°C (Cole et al., 1999), although this impact will be localised.  However, as mussels are of the most damaging biofouling organisms on water outlets of power stations, they are clearly not adversely affected (Whitehouse et al., 1985; Thompson et al., 2000).

Sensitivity assessment. Based on the wide range of temperature tolerance of Mytilus edulis and its limited effect on its physiology, it is concluded that the acute and chronic changes described by the benchmarks of 2-5°C would have limited effect.  Therefore, the biotopes are considered to have a ‘High’ resistance to temperature change, a ‘High’ resilience, and are considered to be 'Not Sensitive' at the benchmark level.

Temperature decrease (local)

Benchmark. A 5°C decrease in temperature for one month, or 2°C for one year. Further detail

Evidence

Local populations may be acclimated to the prevailing temperature regime and may, therefore, exhibit different tolerances to other populations subject to different salinity conditions and therefore caution should be used when inferring tolerances from populations in different regions. Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas which frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek 1992). 

The lower lethal limit of Mytilus edulis depends on the length of time exposed to a low temperature and the frequency of exposure (Bourget, 1983).  Williams (1970) observed that Mytilus edulis tolerated a tissue temperature as low as -10°C.  In a laboratory experiment, Bourget (1983) showed that the median lethal temperature for 24 hours of exposure in Mytilus edulis was -16 °C for large mussels (>3 cm) and -12.5°C for juveniles (<1.5 cm).  However, when exposed to reduced temperatures for only 16 hours, the median lethal temperature of large mussels decreased to -20°C.  It was also reported that mussels exposed to sub lethal temperatures cyclically, e.g. -8°C every 12.4 hours for 3-4 days, suffered significant damage likely to lead to death (Bourget, 1983), which suggested that while Mytilus edulis could tolerate occasional sharp frost events it was not likely to survive prolonged periods of very low temperatures.  During the cold winter of 1962/63, Mytilus edulis was reported to have experienced relatively few effects with only 30% mortality being recorded from the south-east coast of England (Whitstable area) and only about 2% mortality was reported from Rhosilli in South Wales (Crisp, 1964).  Crisp (1964) also noted that the mortality was mainly from predation on the individuals that were weakened by the low temperatures rather than the temperature itself.  It is thought that the use of nucleating agents in the haemolymph and the maintenance of a high osmotic concentration in the mantle fluid during periods of winter isolation allows Mytilus edulis to tolerate such low temperatures (Aunaas et al., 1988).

Shell growth is not expected to be majorly influenced by low temperatures.  Bayne (1976) demonstrated that between 10-20°C water temperature had little effect on the scope for growth, similar to the findings of  (Page & Hubbard, 1987) who found that a temperature range of 10-18°C did not influence growth rate.  In addition, Loo (1992) recorded growth rates of up to 0.7% at temperatures as low as -1°C, with an excess of seston, a rate higher than the same author recorded in mussel culture in Sweden (Loo & Rosenberg, 1983).  They concluded that food availability was more of a limiting factor to growth than temperature (Loo, 1992).

Sensitivity assessment. Based on the wide range of temperature tolerance of Mytilus edulis and its limited effect on its physiology, it is concluded that the acute and chronic changes described by the benchmarks of 2-5°C would have limited effect.  Therefore, the biotopes are considered to have a ‘High’ resistance to temperature change, a ‘High’ resilience, and are considered to be 'Not Sensitive' at the benchmark level.

Salinity increase (local)

Benchmark. A increase in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

Local populations may be acclimated to the prevailing salinity regime and, therefore, exhibit different tolerances to other populations subject to different salinity conditions. Therefore caution should be used when inferring tolerances from populations in different regions.  Mytilus edulis is found in a wide range of salinities from variable salinity areas (18-35 ppt) such as estuaries and intertidal areas to areas of more constant salinity (30-35 ppt) in the sublittoral (Connor et al., 2004).  Furthermore, mussels in rock pools are likely to experience hypersaline conditions on hot days.  Newell (1979) recorded salinities as high as 42 psu in intertidal rock pools, suggesting that Mytilus edulis can tolerate hypersaline conditions.  

Sensitivity assessment.  Increased salinity is likely to change a reduced salinity area to a fully marine area where it is known that mussels can survive in abundance.  Also, an increase in salinity from full to raised salinity (> 40 units), is less than that encountered in rock pools, where Mytilus edulis survives.  Therefore, Mytilus edulis is recorded as having a 'High' resistance to an increase in salinity at the pressure benchmark and a 'High' resilience and this biotope is, therefore, assessed as 'Not sensitive' at the benchmark level.

Salinity decrease (local)

Benchmark. A decrease in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

Local populations may be acclimated to the prevailing salinity regime and may, therefore, exhibit different tolerances to other populations subject to different salinity conditions. Hence, caution should be used when inferring tolerances from populations in different regions. Mytilus edulis is found in a wide range of salinities from variable salinity areas (18-35ppt) such as estuaries and intertidal areas to areas of more constant salinity (30-35 ppt) in the sublittoral (Connor et al., 2004).  In addition, Mytilus edulis thrives in brackish lagoons and estuaries, although, this is probably due to the abundance of food in these environments rather than the salinity (Seed & Suchanek, 1992).Furthermore, mussels in rock pools are likely to experience hypersaline conditions on hot days.  Newell (1979) recorded salinities as high as 42psu in intertidal rock pools, suggesting that Mytilus edulis can tolerate high salinities.  Also, Mytilus edulis was recorded to grow in a dwarf form in the Baltic sea where the average salinity was 6.5psu (Riisgård et al., 2013). 

Mytilus edulis is an osmoconformer and maintains its tissue fluids iso-osmotic (equal ionic strength) with the surrounding medium by mobilisation and adjustment of the tissue fluid concentration of free amino acids (e.g. taurine, glycine and alanine) (Bayne, 1976; Newell, 1989).  But mobilizing amino acids may result in loss of protein, increased nitrogen excretion and reduced growth. However, Koehn (1983) and Koehn & Hilbish (1987) reported a genetic basis to adaptation to salinity.   Mytilus edulis exhibits a defined behavioural response to reducing salinity, initially only closing its siphons to maintain the salinity of the water in its mantle cavity, which allows some gaseous exchange and therefore maintains aerobic metabolism for longer.  If the salinity continues to fall the valves close tightly (Davenport,1979; Rankin & Davenport, 1981).  In the long-term (weeks) Mytilus edulis can acclimate to lower salinities (Almada-Villela, 1984; Seed & Suchanek 1992; Holt et al.,1998).  Almada-Villela (1984) reported that the growth rate of individuals exposed to only 13 psu reduced to almost zero but had recovered to over 80% of control animals within one month.  Observed differences in growth are due to physiological and/or genetic adaptation to salinity.

Decreased salinity has physiological effects on Mytilus edulis; decreasing the heart rate (Bahmet et al., 2005), reducing filtration rates (Riisgård et al., 2013), reducing growth rate (Gruffydd et al., 1984) and reducing the immune function (Bussell et al., 2008).  Both Bahmet et al. (2005) and Riisgård et al. (2013) noted that filtration and heart rates return to normal within a number of days acclimation or a return to the original salinity.  However, Riisgard et al. (2013) observed that mussels from an average of 17 psu found it harder to acclimate between the salinity extremes than those from an average of 6.5 psu.  This observation may mean that mussels in a variable/ lower salinity environment are more able to tolerate change than those found at fully marine salinities.  A sharp salinity change also induces a behavioural response to close the shell (Riisgård et al., 2012) to maintain the salinity within the mantle cavity. In extreme low salinities, e.g. resulting from storm runoff, large numbers of mussels may be killed (Keith Hiscock pers comm.).  However, Bailey et al. (1996) observed very few mortalities when exposing Mytilus edulis to a range of salinities as low as 0ppt for two weeks at a range of temperatures.  It was also noted that there was a fast recovery rate. 

Sensitivity assessment, Most of the literature found on this topic considered short-term (days to weeks) impacts of changes in salinity whilst the benchmark refers to a change for one year.  However, Mytilus edulis was shown to be capable of acclimation to changes in salinity.  As Mytilus edulis is found in salinities to as low as 4-5 psu (Riisgård et al., 2013), it is likely to be able to acclimate to a decrease in salinity from full (30-35 units) to variable (18-40 units) or reduced (18-30 units) . Therefore, Mytilus edulis is recorded as having a ‘High’ resistance to a decrease in salinity and a ‘High’ resilience (no impact to recover from).  The blue mussel bed biotopes are, therefore, considered to be ‘Not Sensitive’ at the benchmark level.    

Water flow (tidal current) changes (local)

Benchmark. A change in peak mean spring bed flow velocity of between 0.1 m/s to 0.2 m/s for more than one year. Further detail

Evidence

Blue mussels are active suspension feeders generating currents by beating cilia and are therefore not entirely dependent on water flow to supply food (organic particulates and phytoplankton).  Therefore, they can survive in very sheltered areas, but water flow (due to tides, currents or wave action) can enhance the supply of food, carried from outside the area or resuspended into the water column.

The growth rate of Mytilus edulis in relation to water flow was investigated by Langan & Howell (1994) who found that the growth rate over 24 days was 0.1, 1.8, 2.0, 1.9 and 1.5 mm at flow rates of 0, 0.01, 0.02, 0.04 and 0.08 m/s respectively.  The only growth rate found to be significantly different was at zero flow.  However, the pattern did follow that predicted by the “inhalant pumping speed” hypothesis that suggested maximal growth at water speeds of about 0.02 m/s and decreased growth rates at higher and lower speeds (Langan & Howell 1994). Higher current speed brings food to the bottom layers of the water column, and hence near to the mussels, at a higher rate (Frechette et al., 1989). Frechette et al. (1989) developed a model based on measurements in the St. Lawrence River estuary (Québec).  The model suggested that Mytilus edulis consumption rate depends on the flow of water.

Widdows et al. (2002) found that there was no change in filtration rate of Mytilus edulis between 0.05 and 0.8 m/s.  They noted that their finding contradicted earlier work that found a marked decline in filtration rates from 0.05 to 0.25 m/s (Newell, 1999; cited in Widdows et al., 2002) but suggested that the difference might be caused in differences in population studied, as the earlier work was based in the USA and their study used mussels from the Exe estuary in the UK.  Widdows et al. (2002) also noted that above 0.8 m/s the filtration rate declined mainly because the mussels became detached from the substratum in the experimental flume tank.  Widdows et al. (2002) noted that their results were consistent with field observations, as mussels show preferential settlement and growth in areas of high flow, such as the mouth of estuaries and at the base of power station cooling systems (Jenner et al., 1998).  They also reported that Jenner et al. (1998; cited in Widdows et al., 2002) observed that biofouling of cooling water systems by mussels was only reduced significantly when mean current speeds reached 1.8-2.2 m/s and mussel biofouling was absent at >2.9 m/s.

Increased flow rate increases the risk of mussels being detached from the bed and transported elsewhere where their chance of survival will be significantly reduced due to the risk of predation and siltation (Dare, 1976).  It is the strength of the byssal attachment that determines the mussel’s ability to withstand increases in flow rate.  Flow rate itself has been shown to influence the strength and number of byssus threads that are produced by Mytilus edulis and other Mytilus spp., with mussels in areas of higher flow rate demonstrating stronger attachment (Dolmer & Svane, 1994; Alfaro, 2006).  Dolmer & Svane (1994) estimated the potential strength of attachment for Mytilus edulis in both still water and flows of 1.94 m/sec, by counting the number of established byssus threads and measuring the strength of attachment of individual detached byssus threads.  It was found that in still water the strength of the attachment was 21% of the potential strength whilst at 19.4 cm/sec it was 81% of the potential strength, suggesting that Mytilus edulis has the ability to adapt the strength of its attachment based on flow rate.  Young (1985) demonstrated that byssus thread production and attachment increased with increasing water agitation.  She observed the strengthening of byssal attachments by 25% within eight hours of a storm commencing and an ability to withstand surges up to 16 m/s.  However, it was concluded that sudden surges may leave the mussels susceptible to being swept away (Young, 1985) as they need time to react to the increased velocity to increase the attachment strength.  Mytilus edulis beds could, therefore, adapt to changes in water flow at the pressure benchmark. 

Alfaro (2006) found that when a sudden increase in flow (to 0.13 m/s) was experienced by Perna canalicuulus (another mussel species) in areas of low flow rate they were more susceptible to detachment than those that had been exposed to a higher flow rate.  It was also noted that the individuals kept at higher water flows (e.g. 10 cm/sec) produced more byssus threads.  The increased energy used for byssus production in the high flow environments may reduce the energy that is available for other biological activities (Alfaro, 2006).

Individuals attached to solid substrata (rock) are likely to display more resistance than individuals attached to boulders, cobbles or sediment.  For example, mussel reefs in the Wash, Morecambe Bay and the Wadden Sea are vulnerable to destruction by storms and tidal surges (Holt et al., 1998).  Widdows et al. (2002) examined mussel beds in the mouth of the Exe estuary and along the coast at Exmouth.  In flume tank studies between 0.1 and 0.35 m/s, the resuspension rate of sediment in mussel beds on sandy substrata was four and five times higher for areas with 25% and 50% mussel cover compared to bare sediment due to the increased turbulence and scouring around the mussels.  However, at high densities (100% cover) the beds remained stable (up to 0.35 m/s), with resuspension being about three times lower than areas with 0% cover, due to the high number of byssal attachments between individuals (Widdows et al., 2002). Where mussel beds occurred on pebble and sand substrata (mixed substrata) sediment erosion was lower than that of the 100% cover on the sandy substrata regardless of mussel density.  Low density mussel beds formed small clumps with a lower mass ratio of mussels attached to the substratum to increase anchorage.  In low density beds, increased scour resulted in some mussel detaching from the bed and in areas with 50% cover the erosion of the bed resulted in the burial of a large proportion of the mussels.  The mussels returned to the surface afterwards and recovered in 1-2 days.  Widdows et al. (2002) also noted a linear relationship between mussel beds density and sediment stability on cohesive mud substratum, taken from Cleethorpes, and exposed to currents of 0.15 to 0.45 m/s.  Again increased mussel cover increased sediment stability.  Widdows et al. (2002) found that the mussel bed at Exmouth experienced a peak flow of 0.9m/s before and after high water, which only reduced to 0.2 m/s at slack water. 

Water flow also affects the settlement behaviour of larvae.  Alfaro (2005) observed that larvae settling in a low water flow environment are able to first settle and then detach and reattach displaying exploratory behaviour before finally settling and strengthening their byssus threads.  However, larvae settling in high flow environments did not display this exploratory behaviour.  Pernet et al. (2003) found that at high velocities, larvae of Mytilus spp. were not able to able to exercise much settlement preference.  It was thought that when contact with suitable substratum is made the larvae probably secure a firm attachment.  Movement of larvae from low shear velocities, where they use their foot to settle, to high shear velocities where they use their byssal thread to settle was observed by Dobretsov & Wahl (2008).

Sensitivity assessment. The blue mussel bed biotopes assessed are recorded from weak (<0.5 m/s) to strong (up to 3 m/s) tidal streams.  The sensitivity of sedimentary biotopes to increased flow is dependent on the substratum and the degree of cover, with dense beds of ca 100% cover being more stable than patchy beds, and more stable on mixed substrata with cobble and boulders than sand and mud.  Connor et al. (2004) noted that the build-up of mussel mud beneath beds could result in a change from sandy to muddy substrata underneath the bed and reduce attachment resulting in increased risk of removal by storms.

A decrease in water flow is unlikely to affect adversely blue mussel beds directly.  The evidence above suggests that they can grow at water flow as low as 0.01 – 0.02 m/s and filter at 0.05 m/s; significantly less than weak tidal streams (<0.5 m/s).  At very low or negligible water flow, the effects of siltation may have adverse effects (see relevant pressure). 

Dense beds are probably stable on mixed substrata, so resistance to change in water flow is probably ‘High’, resilience is assessed as ‘High’ (no effect to recover from) and the biotope is assessed as ‘Not sensitive’.  However, on sandy substrata and possibly to a greater extent on muddy substrata, especially where mussel mud has accumulated and/or the beds are patchy; an increase of water flow at the benchmark level is likely to remove parts of the bed.  Therefore, resistance to change in water flow is probably ‘Medium’, resilience is assessed as ‘Medium’ and the biotopes are assessed as having ‘Medium’ sensitivity. 

Emergence regime changes

Benchmark.  1) A change in the time covered or not covered by the sea for a period of ≥1 year or 2) an increase in relative sea level or decrease in high water level for ≥1 year. Further detail

Evidence

Mytilus edulis beds are found at a wide range of shore heights from in the strandline down to the shallow sublittoral (Connor et al., 2004).  Their upper limits are controlled by temperature and desiccation (Suchanek, 1978; Seed & Suchanek 1992; Holt et al., 1998) while the lower limits are set by predation, competition (Suchanek, 1978) and sand burial (Daly & Mathieson, 1977).  Mussels found higher up the shore display slower growth rates (Buschbaum & Saier, 2001) due to the decrease in time during which they can feed and also a decrease in food availability.  It has been estimated that the point of zero growth occurs at 55% emergence (Baird, 1966) although this figure will vary slightly depending on the conditions of the exposure of the shore (Baird, 1966; Holt et al., 1998). Increasing shore height does, however, increase the longevity of the mussels due to reduced predation pressure (Seed & Suchanek 1992; Holt et al., 1998), resulting in a wider age class of mussels found on the upper shore. The lower limit of Mytilus beds is mainly set by predation from Asterias rubens and Carcinus maenas which may increase with a decrease in emergence potentially reducing the lower limit or reducing the number of size classes and age of the mussels at the lower range of the bed (Saier, 2002). 

Sensitivity assessment. An increase in emergence is likely to reduce the abundance of the biotope at the upper limit due to desiccation and the mussels are likely to be replaced by barnacles on rocky shores, particularly on sheltered shores where there is little spray.  Increased emergence is likely to have more of an impact on sandy substrata where burial by sand could limit the mussels’ ability to extend its lower range resulting in a reduced mussel bed.  A decrease in emergence would probably reduce the abundance of mussels at the lower limit due to increased predation but allow the bed to extend further up the shore. 

Intertidal Mytilus edulis beds are considered to express ‘Medium’ resistance to changes in emergence with mussels at the upper and lower limits exhibiting the greatest effects.  Resilience is assessed as ‘Medium’ and sensitivity is, therefore, assessed as ‘Medium’.  

Wave exposure changes (local)

Benchmark. A change in near shore significant wave height of >3% but <5% for more than one year. Further detail

Evidence

Blue mussel beds are found in a wide range of wave exposures, from extremely exposed areas to extremely sheltered (Seed 1976; Connor et al.,2004).  The littoral mixed sediment biotope (LS.LBR.LMus.Myt.Mx) is found from wave exposed to very sheltered conditions, the sand and mud biotopes (LS.LBR.LMus.Myt.Sa and LS.LBR.LMus.Myt.Mu) occur in moderate wave exposure to sheltered conditions, while LS.LSa.St.MytFab occurs in sheltered conditions (sheltered to extremely sheltered).

Mytilus edulis is able to increase the strength of their attachment to the substratum in more turbulent conditions (Price, 1982; Young, 1985).  Young (1985) demonstrated an increase in strength of the byssal attachment by 25% within 8 hours of a storm commencing.  When comparing mussels in areas of high flow rate and low flow rate those at a higher flow rate exhibit stronger attachments than those in the areas of lower flow (Dolmer & Svane, 1994; Alfaro, 2006).  Dolmer & Svane (1994) found that in still water the strength of the attachment was 21% of the potential strength whilst at 1.94 m/sec it was 81% of the potential strength.  Alfaro (2006) also noted that the individuals kept at higher water flows produce more byssal threads.  The increased energy used for byssus production in the high flow environments may reduce the energy that is available for other biological activities (Alfaro 2006).  Whilst this clearly demonstrates the ability of mussels to adapt to the various conditions to avoid dislodgement, the mussels are unlikely to adapt instantly and a sudden increase in flow is likely to result in dislodgement (Young, 1985). 

Widdows et al. (2002) examined mussel beds in the mouth of the Exe estuary and along the coast at Exmouth.  Where the mussel beds occurred on sandy substratum the re-suspension rate was four and five times higher for areas with 25% and 50% mussel cover compared to bare sediment due to the increased turbulence and scouring around the mussels.  In low density beds, this increased scour resulted in some mussel detaching from the bed and in areas with 50% cover the erosion of the bed resulted in the burial of a large proportion of the mussels.  The mussels returned to the surface after 1-2 days and recovered.  However, at high densities (100% cover) the beds remained stable, with re-suspension being about 3 times lower than areas with  0% cover, due to the high number of byssal attachments between individuals (Widdows et al., 2002). Where mussel beds occurred on pebble and sand substratum (mixed substratum) sediment erosion was lower than that of the 100% cover on the sandy substratum regardless of density despite experiencing flows of 0.9 m/s.  The low density mussels were observed to form small clumps with a lower mass ratio of mussels attached to the substratum to increase anchorage.

Widdows et al. (2002) suggest that 100% mussel cover on sandy substrata reduces the risk of dislodgement.  However, Harger & Landenberger (1971) suggest that growth in mussel beds results in fewer mussels being attached to the substratum and therefore strong seas can “roll up the whole mass of mud and mussels like a carpet and break it to pieces on the foreshore”.  It was also noted that on gravelly substratum, single layer mussel beds incurred less damage in storm conditions than heavier multi-layered beds (Harger & Landenberger,1971). 

Large-scale destruction of mussel beds has been reported in many areas such as the Wash, Morecambe Bay and the Wadden Sea (Holt et al., 1998).  It appears that because of high wave exposure and destruction, reefs found in wave exposed areas are likely to be more dynamic (Nehls & Thiel, 1993).  Furthermore, increased wave exposure leads to a higher risk of damage from drift logs (or other flotsam), which once they have destroyed a patch of mussels, leave the mussels around that patch at a higher risk of erosion (Seed & Suchanek, 1992).  Mussels with high abundance of epizoic and epiphytic (e.g. barnacles and macroalgae) growing on mussels are also more susceptible to removal in areas of high wave exposure due to increased drag caused by these fouling organisms (Suchanek, 1985; Seed & Suchanek, 1992).  However, mussel beds are prevalent in areas of high wave exposure suggesting a high resilience despite the destruction.

Blue mussels display a high resistance to increases in water flow, but the oscillatory water movement that occurs on shores of higher wave exposure is likely to have a higher impact due to the ‘to and fro’ motion which is more likely to weaken the attachments.  Westerbom & Jattu (2006) found that in subtidal mussel beds, mussel densities increased with increasing wave exposure.  The highest biomass was found in areas of intermediate exposure, potentially due to the larger mussels being removed at high wave exposure levels.  It was suggested that the lower densities found in more sheltered areas were due to low recruitment, early post-recruitment mortality, increased predation or stagnant settlement on rocks.  Furthermore, it was also noted that high sedimentation which is more prevalent in sheltered areas, as there is less energy for re-suspension, prevents colonization and result in the death of small mussels that are living close to the sediment surface by smothering and the clogging up of their feeding apparatus (Westerbom & Jattu, 2006). Therefore, colonization of new space in sheltered areas could be slow, particularly in areas where there is a low availability of adult mussels.

An increase in wave exposure may increase density in subtidal beds (Westerbom & Jattu, 2006) unless there is a very sudden storm surge.  Mussels on sedimentary substrata are exposed to a higher risk of dislodgement (Widdows et al., 2002). A decrease in wave exposure is likely to result in increased sedimentation and reduced densities (Westerbom & Jattu, 2006) although the risk of dislodgement will be greatly reduced creating more stable beds (Nehls & Thiel, 1993). 

The above evidence is variable as different studies have examined beds that differ in habitat, wave exposure, substratum and mussel density.  However general trends can be seen.  In rocky habitats, increased wave exposure allows mussel to dominate and form beds, especially where the rock surface has a low slope.  Where the beds are patchy or damaged (from natural or human activities) they are more susceptible to further damage as a result of wave action or storms (Seed & Suchanek, 1992; Brosnan & Crumrine, 1994).  Multi-layered mussel beds are less susceptible to damage, especially where only the surface layer is removed.  It has been noted that the build-up of mussel mud (pseudofaeces) under the bed can reduce the attachment of the bed to the underlying substratum.  But in areas of wave exposure, the flow of water through the bed will probably prevent the ‘mussel mud’ accumulating.

On sedimentary habitats, which themselves occur in wave sheltered environments, the mussel beds stabilise the sediment surface (Widdows et al., 2002), especially at high percentage cover, although at low cover (e.g. in patchy beds) turbulent flow caused by the mussels may increase erosion of the sediment.  Coarse and mixed sediments were more stable, although Widdows et al. (2002) also noted that cohesive muds were also stabilised by mussel beds.  Nevertheless, strong wave action or storms can roll up an entire bed or section of a bed (Harger & Landenberger, 1971), and presumably remove patches of mussels, and that multi-layered bed suffer more damage.  In sedimentary wave sheltered habitats, the build-up of mussel muds may reduce attachment to the substratum and increase the susceptibility of the bed to wave action (Seed & Suchanek, 1992).  The growth of other organisms on the mussels themselves, will increase drag and hence increase the possibility of damage due to wave action.  In sheltered conditions, large macroalgae (e.g. kelps, fucoids) growing on mussels may result in removal of clumps of mussels.

Sensitivity assessment. A decrease in wave exposure is unlikely to adversely affect beds in sheltered, sedimentary habitats, except that muddy sediment will probably increase.  Blue mussel beds on sediment may be more susceptible to damage, as increased wave height increases the possibility of a piece of the bed being removed, or even ‘rolled up, especially in stormy weather.  Therefore, a resistance of ‘Low’ is suggested, with a resilience of ‘Medium’, resulting in a sensitivity of ‘Medium’.  

Transition elements & organo-metal contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Transitional elements & organometal' contaminants on Mytilus spp. are summarized below.  The full 'Mytilus evidence review' should be consulted for details of the studies examined and their results.

In general, the evidence suggested that longer exposure times were required to understand the true impacts of metal exposure on Mytilus, as mussels can close their shells for days.  Hence, short-term exposures (e.g. <48 hrs) may underestimate sensitivity.  This agrees with Widdows & Donkin (1992) who suggested that LC₅₀ values in Mytilus gave a false impression of high tolerance because adult bivalves were able to close their valves and isolate themselves from extreme (potentially lethal) conditions for long periods (i.e. days).  Different life stages had different sensitivities.  This also agrees with Widdows & Donkin (1992) who noted that adults were >10-fold more sensitive than larvae to copper (Cu), petroleum hydrocarbons and sewage sludge.

The majority of the evidence examined copper, followed by cadmium, zinc, silver, and mercury (see Mytilus evidence review; Figure 1.3; Table 1.3).  The evidence suggests that Mytilus adults and juveniles have a ‘High’ sensitivity to copper, cadmium, mercury and silver and a ‘Medium’ sensitivity to iron, lead, methylmercury and neodymium.  The confidence in those assessments is probably ‘Medium’ due to the volume of evidence examined.  However, it is also clear that there is considerable variation in response to metal exposure, due in part to the variation in the experimental studies, and especially the concentration and exposure duration used.

Less evidence for the remaining metals and especially the organometals and nanoparticulate metals was found, and in some cases, the sensitivity assessment is based on one or two papers (e.g. nanoparticulate Zinc, or tributyltin oxide).  While the articles present are all ‘High’ to ‘Medium’ quality and directly applicable, it may be prudent to treat these assessments with more caution and assess their confidence as ‘Low’. 

The number of articles that reported the effects of metals on larvae and embryos alone is also dominated by studies on the effect of copper (Mytilus evidence review; Table 1.4).  The evidence suggests that Mytilus larvae and embryos are highly sensitive to copper, lead, and zinc, plus molybdenum and manganese although the last two are based on single papers.  There is also evidence that organotins result in severe mortality in larvae and embryos. 

Across the entire 'Transitional elements & organometal' contaminant group, there is evidence that several metals, one nanoparticulate metal, and some organometals have been reported to cause ‘severe’ (>75%) mortalities in adult and juvenile mussels (resistance is 'None').  Hence, an overall assessment of ‘High’ sensitivity to 'Transitional elements & organometal' contamination may be given based on the ‘worst-case’ scenario. Overall, the confidence in the assessment is probably ‘Medium’ due to the volume of evidence examined.

Hydrocarbon & PAH contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Hydrocarbons and PAH' contaminants on seagrasses are summarized below. The full 'Mytilus evidence review' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Hydrocarbon' contaminant examined, together with an overall assessment for the pressure. 

Oil spills. Little evidence on the direct physical effects of oil (smothering, or clogging) on Mytilus spp., was found and few studies examined blue mussel beds, except in Babcock et al. (1998) and Rostron & Bunker (1997).  The evidence suggests that Mytilus spp. can be relatively tolerant of direct oiling (in the absence of dispersants or other cleaning treatments) and survived oil spilt by the Torrey Canyon and Sea Empress.  In particular, blue mussel beds in Prince William Sound (Babcock et al., 1998) survived direct oiling and continued exposure to oil retained in the sediment underneath the mussel beds for 3-4 years, although their condition was impaired.  However, Mytilus trossulus abundance in other intertidal habitats was reduced significantly after the Exxon Valdez spill (Highsmith et al., 1996).  In addition, a significant reduction in Mytilus galloprovincialis abundance was also noted after the Hebei Spirit spill in Korea (Jung et al., 2015).  Hence, the effect of oil spills on Mytilus spp. and blue mussel beds is likely dependent on the type of oil spilt, the local habitat, and wave conditions at the time of the spill.  Therefore, resistance is assessed as ‘Low’ to represent the potential for mortality.  Resilience is probably ‘Medium’ so sensitivity to oil spills is assessed as ‘Medium’.

Petroleum hydrocarbons (oils). Refined oils (e.g. lubricant and fuel oils) were reported to be more toxic than crude oils.  Widdows et al. (1982) also noted that the 30-36 µg/l WAF concentrations used in their experiments were comparable to levels found in the environment (e.g. the Thames in 1980) but that very high concentrations (5-1,000 mg/l) were required to elicit a lethal response in Mytilus edulis (see Craddock, 1977).  Overall, the evidence suggests (10% of articles on the effects of oils) that exposure to oils or their water saturated (WSF) or water accommodated fraction (WAF) can result in ‘severe’ mortality (>75%) while another 30% of the articles report significant (25-75%) mortality depending on the type of oil and its concentration.  Therefore, resistance is assessed as ‘None’.  Resilience is probably ‘Low’ so sensitivity to petroleum-based oils is assessed as ‘High’.

Polyaromatic hydrocarbons (PAHs). Only a few articles demonstrated ‘some’ mortality (<25%) due to exposure to PAHs, and then indirectly, as a result of stress and subsequent reduction in the specimen's ability to survive in air.  Similarly, Widdows and others (1995, 2002b) demonstrated a decrease in condition or SFG due to PAH exposure and body burden.  However, most articles examined (93%) only reported sub-lethal effects.  Therefore, resistance is assessed as ‘Medium’ to represent the ‘worst-case’ potential of PAHs to cause indirect mortality due to reduced condition and/or stress.  Resilience is probably ‘Medium’ so sensitivity to PAHs is assessed as ‘Medium’.

Others. The evidence on ‘other’ forms of hydrocarbons was limited.  Toluene is potentially toxic to Mytilus spp. (Sabourin & Tullis,1981), while benzene, olive oil mill wastewater, styrene and ‘tetralins’ were reported to have sub-lethal effects at the concentrations studied.

Sensitivity to 'Hydrocarbons and PAH' contamination. In their review, Widdows & Donkin (1992) note that (one reason) mussels are good sentinels for pollution is because they are relatively tolerant of, but not insensitive, to a range of environmental conditions and contaminants. Furthermore, they noted that adults were >10-fold more sensitive than larvae to copper (Cu), petroleum hydrocarbons and sewage sludge. Widdows & Donkin (1992) noted that lethal responses give a false impression of high tolerance since the adults can close their valves and isolate themselves from the environment for days. They suggested that sub-lethal effects e.g., shell growth and 'scope for growth' (SFG), were more sensitive indicators of the effects of contaminants.

The evidence review suggests that exposure to hydrocarbon contamination can cause mortality in Mytilus spp., which is in some cases 'significant' or even 'severe'.  The degree of mortality, or absence of mortality, depends on the type of hydrocarbon (crude or refined oils, oil-saturated water fractions, PAHs, or refined products) to which the species is exposed, how they are exposed (through oil spills, effluents, the sediment, or food supply e.g. algae), the concentration of the contaminant and the duration of exposure, as well as seasonal influences on the species’ condition, especially spawning and reproduction.

Therefore, the 'weight of evidence' based on reported 'severe' (>75%) and 'significant' (25-75%) mortality due to hydrocarbon contamination suggests an overall ‘worst case’ resistance assessment of ‘None’.  Resilience is probably ‘Low’ so sensitivity to petroleum-based oils is assessed as ‘High’.  However, it should be noted that the evidence reviewed also documented several occasions in which blue mussels and blue mussel beds had survived significant oiling and most evidence (70% of the articles examined) of exposure to hydrocarbons was reported to result in sub-lethal effects, although it was not clear how detrimental sub-lethal effects or ‘stress’ is to the species survival. Hence, confidence in the assessments is ‘Medium’.

Synthetic compound contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Synthetic compound' contaminants on Mytilus spp. are summarized below. The full 'Mytilus evidence review' should be consulted for details of the studies examined and their results. 

In general, the evidence suggested that longer exposure times were required to understand the effects of exposure to synthetic contaminants on Mytilus, as mussels could close their shells for days.  Hence, short-term exposures (e.g. <48hrs) may underestimate sensitivity.  This agrees with Widdows & Donkin (1992) who suggested that LC₅₀ values in Mytilus gave a false impression of high tolerance because adult bivalves were able to close their valves and isolate themselves from extreme (potentially lethal) conditions for long periods (i.e. days).

The majority of articles reported a lethal response of exposure to synthetic compounds in Mytilus spp.  A total of 57% of ranked mortalities reported in the evidence review were lethal (‘Severe’, ‘Significant’ or ‘Some’), while 27% reported no mortality (‘None’) and 16% reported sub-lethal effects. Most of the articles examined pesticides/biocides and pharmaceuticals (Mytilus evidence review; Figure 1.10).  A total of 15 (56%) of the 27 articles that examined pesticides reported lethal effects.  The majority of the evidence suggested that pesticides resulted in lethal effects in adults and juvenile Mytilus spp. but that larval and embryos were probably more sensitive.  Therefore, we can suggest that Mytilus spp. probably has a ‘High’ sensitivity to pesticide exposure, with a few exceptions.  The confidence in the assessment is assessed as ‘Medium’ because of the number of articles examined and the consistency in the response.

However, 19 (70%) of the articles that examined pharmaceuticals reported lethal effects.  The most lethal responses were shown by the larvae and embryos rather than adults and juveniles.  Therefore, we can suggest that Mytilus spp. probably has a ‘High’ sensitivity to the pharmaceuticals examined especially in the larvae and developmental stages.  The confidence in the assessment is assessed as ‘Medium’ because of the number of articles examined and the consistency in the response.

The evidence on other synthetic contaminant types is more limited.  The flame retardant Tetrabromo bisphenol A (TBBPA) caused mortality and abnormal development in larvae (Fabbri et al., 2014) while another two flame retardants had no significant effects on adults (Barón et al., 2016).  Different types of surfactant caused lethal responses in larvae, embryos and in adults.  PFAS exposure caused mortality in larvae and embryos but no studies on the effects on adults were found. 

Nevertheless, the results (Mytilus evidence review; Table 1.7 & 1.8) suggest that Mytilus spp. are probably sensitive to a number of synthetic compounds, especially in early development or as larvae.  Therefore, the sensitivity of Mytilus spp. to the ‘Synthetic compounds’ examined is assessed as ‘High’ (resistance is 'None' and resilience is 'Low'), especially in larvae and developmental stages.  Overall, the confidence in the assessment is probably ‘Medium’ because of the number of articles examined and the consistency in the response.

Radionuclide contamination

Benchmark. An increase in 10µGy/h above background levels. Further detail

Evidence

The periostracum of Mytilus edulis was reported to concentrate uranium (Widdows & Donkin, 1992). Mussels have also been reported to bioaccumulate ¹⁰⁶Ru, ⁹⁵Zr, ⁹⁵Nb, ¹³⁷Cs and ⁹⁰Sr (Cole et al., 1999). While the above data demonstrates that Mytilus edulis can accumulate radionuclides, little information concerning the effects of radionuclides on marine organisms was found. Sensitivity to this pressure is therefore not assessed based on lack of evidence. 

Introduction of other substances

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

No evidence of the effects of the 'Introduction of other substances' (see pressure definition) was found.

De-oxygenation

Benchmark. Exposure to dissolved oxygen concentration of less than or equal to 2 mg/l for one week (a change from WFD poor status to bad status). Further detail

Evidence

Mytilus edulis is capable of anaerobic metabolism.  In aerial exposure (emersion) the mussel closes its valves, resulting in a low rate of oxygen exchange and consumption, and conservation of energy (Widdows et al., 1979a; Zwaan de & Mathieu 1992). 

Mytilus edulis is regarded as euryoxic, tolerant of a wide range of oxygen concentrations including zero (Zandee et al., 1986; Wang & Widdows 1991; Gosling, 1992; Zwaan de & Mathieu 1992; Diaz & Rosenberg 1995; Gray et al., 2002).  Diaz & Rosenberg (1995) suggest it is resistant to severe hypoxia.   Adult mytilids exhibited high tolerance of anoxia in laboratory tests, e.g. Theede et al., (1969) reported LD₅₀ of 35 days for Mytilus edulis exposed to 0.21 mg/l O₂ at 10°C, which was reduced to 25 days with the addition of sulphide (50 mg/l Na₂S.9H₂O).  Jorgensen (1980) observed, by diving, the effects of hypoxia (0.2 -1 mg/l) on benthic macrofauna in marine areas in Sweden over a 3-4 week period.  Mussels were observed to close their shell valves in response to hypoxia and survived for 1-2 weeks before dying (Cole et al., 1999; Jorgensen, 1980). 

All life stages show high levels of tolerance to low oxygen levels.  Mytilus edulis larvae, for example, are tolerant down to 1.0 ml/l, and although the growth of late-stage larvae is depressed in hypoxic condition, the settlement behaviour does not seem to be affected (Diaz & Rosenberg, 1995).  Based on the available evidence Mytilus edulis are considered to be resistant to periods of hypoxia and anoxia although sub-lethal effects on feeding and growth may be expected.

Sensitivity assessment Mytilus edulis is considered to be not sensitive to de-oxygenation at the pressure benchmark. Resistance is, therefore, assessed as ‘High’, resilience as ‘High’ (no effect to recover from), and sensitivity assessed as 'Not sensitive' at the benchmark level.  

Nutrient enrichment

Benchmark. Compliance with WFD criteria for good status. Further detail

Evidence

This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations. Nutrient enrichment may impact mussel beds by altering the biomass of phytoplankton and macroalgae.  At low levels, nutrient enrichment may stimulate the growth of phytoplankton used as food - a potentially beneficial effect.  In the Wadden Sea, where fishing had caused the destruction of the local population of Sabellaria spinulosa, Mytilus edulis was able to colonize, partly because of the increase in coastal eutrophication (Maddock, 2008).  Conversely, Dinesen et al. (2011) observed that a reduction in nutrient loading to comply with the WFD resulted in a decrease of mussel biomass in estuaries.

High levels of enrichment may stimulate algal blooms and macroalgal growth.  The growth of macrophytes on the mussel beds may result in increased drag on the mussel bed and hence increase susceptibility to damage from wave action and/or storms (see changes in wave exposure pressure).  Algal blooms may die off suddenly, causing de-oxygenation (see de-oxygenation pressure) where the algae decompose on the seabed.  The thresholds at which these blooms occur depend on site-specific conditions and be mitigated by the degree of mixing and tidal exchange. Some algae have been shown to negatively affect Mytilus edulis when present in high concentrations.  For example, blooms of the algae Phaeocystis sp., have been observed to block the mussel's gills when present in high concentrations reducing clearing rates, and at high levels, they caused a complete cessation of clearance (Smaal & Twisk, 1997).  Blockage of the gills is also likely to reduce ingestion rates, prevent growth and cause reproductive failure (Holt et al., 1998).  Other species known to negatively impact Mytilus edulis are Gyrodinium aureolum (Tangen, 1977; Widdows et al., 1979b) and a non-flagellated chrysophycean alga (Tracey, 1988). The accumulation of toxins from algal blooms has also been linked to outbreaks of paralytic shellfish poisoning resulting in the closure of shellfish beds (Shumway, 1990).

Sensitivity assessment. Mytilus edulis beds are considered to be 'Not sensitive' to nutrient enrichment at levels that comply with the requirements for good status for transitional and coastal water bodies (UKTAG, 2014). 

Organic enrichment

Benchmark. A deposit of 100 gC/m²/yr. Further detail

Evidence

Organic enrichment can result from inputs of additional organic matter.  Organic enrichment may lead to eutrophication with adverse environmental effects including deoxygenation, algal blooms and changes in community structure (see nutrient enrichment and de-oxygenation). 

It has been shown that regardless of the concentration of organic matter, Mytilus edulis will maintain its feeding rate by compensating with changes to filtration rate, clearance rates, production of pseudofaeces and absorption efficiencies (Tracey, 1988; Bayne et al., 1993; Hawkins et al., 1996).  A number of studies have highlighted the ability of Mytilus edulis to utilise the increased volume of organic material available at locations around salmon farms.  Reid et al. (2010) noted that Mytilus edulis could absorb organic waste products from a salmon farm with great efficiency.  Increased shell length, wet meat weight, and condition index were shown at locations within 200 m from a farm in the Bay of Fundy allowing a reduced time to market (Lander et al., 2012). Mytilus edulis have also been recorded in areas around sewage outflows (Akaishi et al., 2007; Lindahl & Kollberg, 2008; Nenonen et al., 2008; Giltrap et al., 2013) suggesting that they are highly tolerant of the increase in organic material that would occur in these areas. It should be noted that biotopes occurring in tide-swept or wave exposed areas are less likely to experience the effects of organic enrichment as the organic matter will be rapidly removed.

Sensitivity assessment. Based on the observation of Mytilus edulis thriving in areas of increased organic matter (Lander et al., 2012, Reid et al., 2010), it was assumed that Mytilus edulis beds had a ’High’ resistance to increased organic matter at the pressure benchmark.  Resilience is, therefore, assessed as ‘High’ (no effect to recover from) and sensitivity assessed as 'Not sensitive'. 

Physical loss (to land or freshwater habitat)

Benchmark. A permanent loss of existing saline habitat within the site. Further detail

Evidence

All marine habitats and benthic species are considered to have a resistance of ‘None’ to this pressure and to be unable to recover from a permanent loss of habitat (resilience is ‘Very Low’).  Sensitivity within the direct spatial footprint of this pressure is, therefore ‘High’.  Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  Adjacent habitats and species populations may be indirectly affected where meta-population dynamics and trophic networks are disrupted and where the flow of resources e.g. sediments, prey items, loss of nursery habitat etc. is altered.

Physical change (to another seabed type)

Benchmark. Permanent change from sedimentary or soft rock substrata to hard rock or artificial substrata or vice-versa. Further detail

Evidence

Mytilus edulis can be found on a wide range of substrata including artificial substratum (e.g. metal, wood, concrete), bedrock, biogenic reef, caves, crevices / fissures, large to very large boulders, mixed, muddy gravel, muddy sand, rock pools, sandy mud, small boulders, under boulders (Connor et al., 2004). An increase in the availability of hard substratum may be beneficial in areas where sedimentary habitats were previously unsuitable for colonization e.g. coarse, mobile sediments.  It should be noted that differences in diversity and other structural characteristics of assemblages between natural and artificial substratum have been observed suggesting that there is not a direct, compensatory effect. 

Mussels themselves will often cause a change in substrata by the deposition of large quantities of ‘mussel-mud’ composed of faecal matter and pseudofaeces (Dare, 1976) particularly in areas of low water movement.  A change from rock to sand in an area of high water flow would increase the mussels’ vulnerability to dislodgement and scour (Widdows et al., 2002) and, potentially, smothering as sand smothering has been shown to set the lower limit of Mytilus beds in some areas (Daly & Mathieson 1977). 

Sensitivity assessment.  A change in substratum type would not necessarily reduce habitat quality for individual Mytilus edulis which can colonize a wide range of hard and sedimentary habitats. However, a change in substratum type would alter biotope classification. Hence, the resistance of the biotope is assessed as ‘None’ (loss of >75% of extent), resilience (following habitat recovery) is assessed as ‘Very low’ (the pressure is a permanent change).  Sensitivity, based on combined resistance and resilience is assessed as ‘High’. The more precautionary assessment for the biotope, rather than the species, is presented in the table as it is considered that any change to a reef habitat from a sedimentary habitat would alter the biotope classification and hence the more sensitive assessment is appropriate.

Physical change (to another sediment type)

Benchmark. Permanent change in one Folk class (based on UK SeaMap simplified classification). Further detail

Evidence

Mytilus edulis can be found on a wide range of sediment substrata including, mixed, muddy gravel, muddy sand, rock pools, sandy mud, small boulders, under boulders (Connor et al., 2004). It should be noted that the mussels themselves will often cause a change in substrata by the deposition of large quantities of ‘mussel-mud’ composed of faecal matter and pseudofaeces (Dare, 1976) particularly in areas of low water movement.  A change from rock to sand in an area of high water flow would increase the mussels’ vulnerability to dislodgement and scour (Widdows et al., 2002) and, potentially, smothering as sand smothering has been shown to set the lower limit of Mytilus beds in some areas (Daly & Mathieson, 1977). 

Sensitivity assessment The pressure benchmark refers to the simplified Folk classification developed by Long (2006) and the UK Marine Habitat Classification Littoral and Sublittoral Sediment Matrices (Connor et al., 2004).   In most instances, the pathway (human activity) by which the substratum is changed would remove or smother mussels.  However, these effects are addressed under the ‘hydrological change’, ‘abrasion’; ‘penetration and disturbance’ and ‘smothering’ pressures above. 

The natural modification of the sediment due to the build-up of mussel-mud could result in a change of sediment type from mixed and sand dominated, to anoxic mud, depending on location and hydrography. In addition, a change in sediment type would also change the biotope definition.  That is, a change from mud to sand would change the biotope from LS.LBR.LMus.Myt.Mu to LS.LBR.LMus.Myt.Sa.  However, the biotopes LS.LBR.LMus and LS.LBR.LMus.Myt encompasse three variant biotopes found on mud, sand and mixed sediment.  These biotopes encompass the full variety of sediments ranging from mud and sand to pebble, cobbles and medium boulders; although gravel is not mentioned (Connor et al., 2004).  The higher level biotope is, therefore, considered to have ‘High’ resistance and subsequently ‘High' recovery and ‘Not sensitive' to a change in sediment type of 1 Folk class as a change between mixed sediments, mud and sandy mud and sand and muddy sand would not adversely affect the biotopes.  

Habitat structure changes - removal of substratum (extraction)

Benchmark. The extraction of substratum to 30 cm (where substratum includes sediments and soft rock but excludes hard bedrock). Further detail

Evidence

The process of extraction will remove the entire mussel bed and the associated community; therefore a resistance of ‘None’ is recorded.  Hence, resilience is assessed as ‘Low’, and sensitivity as ‘High’.  

Abrasion / disturbance of the surface of the substratum or seabed

Benchmark. Damage to surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

Mytilus edulis lives on the surface of the seabed held by byssus threads attached to either the substratum or to other mussels in the bed.  Activities resulting in abrasion and disturbance can either directly affect the mussel by crushing them, or indirectly affect them by the weakening or breaking of their byssus threads making them vulnerable to displacement (Denny, 1987) where they are unlikely to survive (Dare, 1976).  In addition, abrasion and sub-surface damage may attract mobile scavengers and predators including fish, crabs, and starfish to feed on exposed, dead and damaged individuals and discards (Kaiser & Spencer, 1994; Ramsay et al., 1998; Groenewold & Fonds, 2000; Bergmann et al., 2002).  This effect will increase predation pressure on surviving damaged and intact Mytilus edulis.  A number of activities or events that result in abrasion and disturbance and their impacts on mussel beds are described below, based on the review by Mainwaring et al. (2014). 

The effects of trampling have been more widely studied in the terrestrial community showing that when areas are intensively trampled bare patches are likely to result as a result of erosion or equally soil compaction may result (Liddle, 1997).  There are a number of studies which have focused on the impact of trampling on the intertidal rocky shore whereas the impact on sedimentary shores is relatively poorly studied (Tyler-Walters & Arnold, 2008).  In general, studies have found that trampling is an additional disturbance to the natural disturbances that the intertidal organisms are adapted to tolerate. Large declines of the Mytilus californianus from mussel beds due to trampling have been reported (Brosnan, 1993; Brosnan & Crumrine, 1994; Smith & Murray, 2005).  Brosnan & Crumrine (1994) recorded the loss of 54% of mussels from a single experimental plot on one day.  Mussels continued to be lost throughout the experimental period, forming empty patches larger than the experimental plots.  The empty patches continued to expand after trampling had ceased, due to wave action.  At another site, the mussel bed was composed of two layers, so that while mussels were lost, cover remained.  Brosnan (1993) also reported a 40% loss of mussels from mussel beds after three months of trampling and a 50 % loss within a year.  Van de Werfhorst & Pearse (2007) examined M. californianus abundance at sites with differing levels of trampling disturbance.  The highest percentage of mussel cover was found at the undisturbed site while the severely disturbed site showed low mussel cover. 

Smith & Murray (2005) examined the effects of low-level disturbance on an extensive bed of Mytilus californianus (composed of a single layer of mussels) in southern California.  Smith & Murray (2005) reported that in experimental plots exposed to trampling, mussel loss was 20-40% greater than in untreated plots.  A decrease in mussel mass, density, cover and maximum shell length were recorded even in low intensity trampling events (429 steps/m²).  However, only 15% of mussel loss was as a direct result of trampling, with the remaining loss occurring during intervals between treatment applications. Brosnan & Crumrine (1994) suggested that trampling destabilizes the mussel bed, making it more susceptible to wave action, especially in winter.  Smith & Murray (2005) suggested that an indirect effect of trampling was weakening of byssal threads, which increases mussel susceptibility to wave disturbance (Denny, 1987).  Brosnan & Crumrine (1994) observed recruitment within experimental plots did not occur until after trampling had ceased, and no recovery had occurred within 2 years 

Brosnan and Crumrine (1994) noted that mussels that occupied hard substrata but did not form beds were also adversely affected.  Although only at low abundance (2.5% cover), all mussels were removed by trampling within 4 months.  Brosnan & Crumrine (1994) noted that mussels were not common and confined to crevices in heavily trampled sites.  Similarly, the mussel bed infauna (e.g. barnacles) was adversely affected and were crushed or lost with the mussels to which they were attached.  However, Beauchamp & Gowing (1982) did not observe any differences in mussel density between sites that differed in visitor use.

Paine & Levine (1981) examined natural patch dynamics in a Mytilus californianus bed in the USA.  They suggested that it may take up to seven years for large barren patches to recover.  However, chronic trampling may prevent recovery altogether.  This would result in a shift from a mussel dominated habitat to one dominated by an algal turf or crust (Brosnan & Cumrine, 1994), completely changing the biotope.  However, a small period of trampling could allow communities to recover at a similar rate to that of natural disturbance as the effects are similar.  The associated epifauna and epiflora suffer the greatest amount of damage as they are the first organisms that a foot makes contact with (Brosnan & Crumrine, 1994).  The loss of epifauna and epiflora could initially be of benefit to the mussel bed, despite the obvious decrease in species diversity, as there will be a decrease in drag for the mussels reducing the risk of dislodgement (Witman & Suchanek 1984) and freeing up more energy for growth and reproduction.  However, it is likely that after continued trampling this effect will be minimal compared with the increased risk of dislodgement caused by trampling. No studies assessing the effect of trampling on mussels on intertidal muddy sand or sediments were found.  Losses to the adult mussels by crushing or by suffocation where these are forced into the sediment are expected.  There is the potential that this will open up areas for new recruitment or it may just create a similar situation to that seen on the rocky shore where wave damage and continual trampling prevent settlement and recovery.

The collision of objects with the bed, such as wave driven logs (or similar flotsam), is known to cause the removal of patches of mussels from mussel beds (Seed & Suchanek, 1992; Holt et al., 1998).  When patches occur in mussel beds a good recruitment could result in a rapid recovery or the patch may increase in size through the weakening of the byssus threads of the remaining mussels leaving them vulnerable to erosion from storm damage (Denny, 1987). Damage in areas of high wave exposure is likely to result in increased erosion and a patchy distribution although recruitment may be high.  In sheltered areas, damage may take a lot longer due to limited larval supply, although the frequency of destruction through wave driven logs would be less than in high wave exposure.  Similar effects could be observed through the grounding of a vessel, the dropping of an anchor or the laying of a cable, although the scale of damage clearly differs. Shifting sand is known to limit the range of Mytilus edulis through burial and abrasion (Daly & Mathieson, 1977).

Various fishing methods also result in abrasion of the mussel beds.  Bait collection through raking will cause surface abrasion and the removal of patches of mussel resulting in the damage and recovery times described above.  Holt et al., (1998) reported that hand collection, or using simple hand tools occurs in small artisanal fisheries.  They suggested that moderate levels of collection by experienced fishermen may not adversely affect the biodiversity of the bed.  But they also noted that even artisanal hand fisheries can deplete the mussel biomass on accessible beds in the absence of adequate recruitment of mussels. Smith & Murray (2005) observed a significant decrease in mussel mass (g/m²), density (no./m2), percentage cover and mean shell length due to low-intensity simulated bait-removal treatments (2 mussels/month) for 12 months (Smith & Murray, 2005).  They also stated that the initial effects of removal were ‘overshadowed’ by the loss of additional mussels during time periods between treatments, probably due to the indirect effect of the weakening of byssal threads attachments between the mussel leaving them more susceptible to wave action (Smith & Murray, 2005).  The low-intensity simulated bait-removal treatments had reduced percentage cover by 57.5% at the end of the 12 month experimental period.  Smith & Murray (2005) suggested that the losses incurred from collection and trampling are far greater than those that occur by natural causes.  This conclusion was reached due to significant results being displayed for human impact despite the experiment taking place during a time of high natural disturbance from El Niño–Southern Oscillation (ENSO).

Sensitivity assessment. Based on the available evidence it is concluded that all mussel biotopes are sensitive to abrasion and that resistance is ‘Low’ (loss of 25-75% of bed within direct impact footprint), resilience is assessed as 'Medium’, resulting in a sensitivity of ‘Medium’.

Penetration or disturbance of the substratum subsurface

Benchmark. Damage to sub-surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

Mytilus edulis lives on the surface of the seabed held in one place by byssus threads that either attach to the substratum or to other mussels in the bed.  Activities resulting in penetration and disturbance can either directly affect the mussel by crushing or removal, or indirectly affect them by the weakening or breaking of their byssus threads making them vulnerable to displacement (Denny, 1987) where they are unlikely to survive (Dare, 1976). Where mussels are removed the associated fauna and flora will also be removed.  In addition, abrasion and sub-surface damage attract mobile scavengers and predators including fish, crabs, and starfish to feed on exposed, dead and damaged individuals and discards (Kaiser & Spencer, 1994; Ramsay et al., 1998; Groenewold & Fonds, 2000; Bergmann et al., 2002).  This effect could increase predation pressure on surviving damaged and intact Mytilus edulis. 

Mussel dredging is the main form of activity that results in penetration around mussel beds.  Holt et al. (1998) noted that several thousand tonnes of mussels were fished in the Wash by dredgers in good years.  Dredging will remove the substratum along with the mussels and their associated flora and fauna.  Temporary re-suspension of sediment also occurs with mussel dredging (Holt et al., 1998) in volumes of 1470 g/m² (Riemann & Hoffmann, 1991), which could potentially result in localised smothering.   Dredging is also likely to increase the vulnerability of the remaining mussels to storm damage through the weakening of byssal attachment and creating patches in the bed (Denny, 1987).

The Scottish MPA Project Fisheries Management Guidance (JNCC, 2013a) suggests that scallop dredges and other demersal towed gear are also likely to result in the removal of a proportion of the bed along with its associated fauna and flora.  The same report suggested that potting and other demersal static gear would have a lower impact than mobile gear.  There is no evidence for the impacts of hydraulic dredging on mussels but Hall et al. (1990) observed that when using hydraulic dredging for Ensis sp. the immediate effects were a reduction in the number of target species and many macrofaunal species.  However, after 40 days the effect of the fishing gear could no longer be seen. 

Holt et al. (1998) noted that natural ‘wild’ beds are susceptible over-exploitation, especially in some embayments and that over-exploitation can reduce recruitment.  Holt et al. (1998) also point out that the source areas for recruitment to beds is unknown and the relationship between stock and recruitment poorly understood.  This statement is consistent with the sporadic and unpredictable nature of recruitment and recovery in mussels beds (Seed & Suchanek, 1992).

Sensitivity assessment. The activities that penetrate the seabed could result in removal of part of a bed and its associated fauna and flora. Based on the available evidence it is concluded that all sedimentary mussel biotopes are sensitive to ‘penetration and/or disturbance of the seabed’. Therefore, resistance is assessed as ‘Low’ (loss of 25-75% of bed within direct footprint), resilience is assessed as 'Medium’, and sensitivity as ‘Medium’.  

Changes in suspended solids (water clarity)

Benchmark. A change in one rank on the WFD (Water Framework Directive) scale e.g. from clear to intermediate for one year. Further detail

Evidence

Mytilus edulis does not rely on light penetration for photosynthesis. In addition, visual perception is limited and the species does not rely on sight to locate food or other resources.  An indirect effect of increased turbidity and reduced light penetration may be reduced phytoplankton productivity which could reduce the food availability for Mytilus edulis.  However, as Mytilus edulis uses a variety of food sources and food is brought in from other areas with currents and tides, the effect is likely to be minimal.  This species and the biotopes it forms are therefore not sensitive to changes in water clarity that alter light penetration.

Mytilus edulis is often found in areas with high levels of turbidity.  For example, the average suspended particulate matter (SPM) concentration at Hastings Shingle Bank was 15 -20 mg/l in June 2005, reaching 50 mg/l in windier (force 4) conditions, although a concentration of 200 mg/l was recorded at this site during gales (Last et al., 2011).  Winter (1972, cited by Moore, 1977) recorded 75% mortality of Mytilus edulis in concentrations of 1.84-7.36 mg/l when food was also available.  However, a relatively small increase in SPM concentration e.g. from 10 mg/l to 90 mg/l was found to increase growth rates (Hawkins et al., 1996).  Concentrations above 250 mg/l have been shown to impair the growth of filter-feeding organisms (Essink, 1999).  But Purchon (1937) found that concentrations of particulates as high a 440 mg/l did not affect Mytilus edulis and that mortality only occurred when mud was added to the experiment bringing the concentrations up to 1220 mg/l.  The reason for some of the discrepancy between studies may be due to the volume of water used in the experiment.  Loosanoff (1962) found that in small quantities of turbid water (due to particulates) the mussel can filter out all of the particulates within a few minutes whereas in volumes >50 gallons per individual the mussel becomes exhausted before the turbidity has been significantly lowered, causing it to close its shell and die.

It may be possible for Mytilus edulis to adapt to a permanent increase in SPM by decreasing their gill size and increasing their palp size in areas of high turbidity (Theisen, 1982; Essink, 1999).  In areas of variable SPM, it is likely that the gill size would remain the same but the palp would adapt (Essink, 1999).  Whilst the ability to adapt may prevent immediate declines in health, the energetic costs of these adaptations may result in reduced fitness; the extent of which is still to be established.

Mytilus edulis uses the circadian clock to determine the opening of the shell gape in nocturnal gape cycles (Ameyaw-Akumfi & Naylor, 1987).  Last et al. (2011) investigated the effects on increased SPM concentrations on both the gape pattern and mortality in order to establish the effect that aggregate dredging will have on Mytilus edulis and other benthic invertebrates.  Therefore they tested concentrations similar to those expected within a few hundred meters of an aggregate extraction site.  The highest concentration tested using a pVORT (paddle VOrtex Resuspension Tanks) was ~71 mg/l.  They showed that there is a significant reduction of the strength of the nocturnal gape cycle at high suspended sediment loads as well as a change in the gape period.  The effects of these changes are not fully known but as it is likely that the gape pattern is a strategy to avoid diurnal predators the change may result in an increased risk of predation.  On the other hand, the increased turbidity may reduce predation by visual predators such as fish and birds (Essink, 1999).  After continued measurements of the gape cycle for 4 days post-treatment, Last et al. (2011) observed that the cycle took longer than this to recover from the cycle disruption.  Further study is required to determine the length of time required for recovery of this behavioural response (Last et al., 2011).

Based on a comprehensive literature review, Moore (1977) concluded that Mytilus edulis displayed a higher tolerance to high SPM concentrations than many other bivalves although the upper limit of this tolerance was not certain.  He also hypothesised that the ability of the mussel to clean its shell in such conditions played a vital role in its success along with its pseudofaecal expulsion.  A reduction in SPM concentrations may be caused by the erecting of dams and hydroelectric power stations (Moore, 1977), which could leave subtidal mussel beds more vulnerable to visual predators such as birds and fish.  The recovery time from increased predation pressures would depend on the duration of the reduced turbidity.  If reduced SPM concentration is also linked with a reduction of suspended organic matter then it could be assumed that the mussel fitness would be negatively affected by a reduction in food supply.  However, as active filter feeders, they are not dependent on water flow to supply food.

Sensitivity assessment. Evidence indicates that Mytilus edulis and hence blue mussel beds can tolerate a broad range of suspended solids.  The benchmark for this pressure refers to a change in turbidity of one rank on the Water Framework Directive (WFD) scale.  Mussel beds form in relatively clear waters of open coasts and wave exposed shores, and on sediments in the sheltered coast (where turbulent water flow over the mussel beds could resuspend sediments locally) and in turbid bays and estuaries.  Therefore, is unlikely that a change in turbidity by of one rank (e.g. from 300 to 100 mg/l or <10 to 100 mg/l) will significantly affect the mussel bed.   Hence, resistance to this pressure is assessed as ‘High', recovery is assessed as ‘High’ (no impact to recover from), and sensitivity is assessed as 'Not sensitive' at the benchmark level.  

Smothering and siltation rate changes (light)

Benchmark. ‘Light’ deposition of up to 5 cm of fine material added to the seabed in a single discrete event. Further detail

Evidence

The main human activity that increases sedimentation is dredging and the dumping of dredged sediments in estuarine and coastal waters.  Aggregate dredging and fishing gear can cause localised sedimentation and smothering.  However, changes in water flow can cause localised smothering within mussel beds (Widdows et al., 2002), and storms may move large volumes of sediment and smother entire mussel beds (Dare, 1976). Mytilus edulis occurs in areas of high suspended particulate matter (SPM) and therefore a level of siltation is expected from the settling of SPM.  In addition, the high rate of faecal and pseudofaecal matter production by the mussels naturally results in siltation of the seabed, often resulting in the formation of large mounds beneath the mussel bed.  For example, at Morecambe Bay an accumulation of mussel-mud (faeces, pseudofaeces and washed sand) of 0.4-0.5m between May 1968 and September 1971 resulted in the mortality of young mussels (Daly & Mathieson, 1977).  In order to survive the mussels needed to keep moving upwards to stay on the surface.  Many individuals did not make it to the surface and were smothered by the accumulation of mussel-mud (Daly & Mathieson, 1977), so that whilst Mytilus edulis does have the capacity to vertically migrate through sediment some individuals will not survive.  

Sand burial has been shown to determine the lower limit of Mytilus edulis beds (Daly & Mathieson, 1977a).  Burial of Mytilus edulis beds by large scale movements of sand, and resultant mortalities have been reported from Morecambe Bay, the Cumbrian coast and Solway Firth (Holt et al., 1998).  Essink (1999) recorded fatal burial depths of 1-2 cm for Mytilus edulis and suggested that they had a low tolerance of sedimentation based on investigations by R.Bijkerk (cited by Essink, 1999).  Essink (1999) suggested that deposition of sediment (mud or sand) on shallow mussel beds should be avoided.  However, Widdows et al. (2002) noted that mussels buried by 6 cm of sandy sediment (caused by resuspension of sediment due to turbulent flow across the bed) were able to move to the surface within one day.  Conversely, Condie (2009) (cited by Last et al., 2011) reported that Mytilus edulis was tolerant of repeated burial events. 

Last et al., (2011) carried out burial experiments on Mytilus edulis in pVORTs.  They used a range of burial depths and sediment fractions and temperatures.  It was found that individual mussels were able to survive burial in depths of 2, 5 and 7 cm for over 32 days although the deeper and longer the mussels were buried the higher the mortality.  Only 16 % of buried mussels died after 16 days compared to almost 50 % mortality at 32 days.  Mortality also increased sharply with a decrease in particle size and with increases in temperature from 8.0 and 14.5 to 20 °C.  The ability of a proportion of individuals to emerge from burial was again demonstrated with approximately one quarter of the individuals buried at 2cm resurfacing.  However, at depths of 5 cm and 7 cm no emergence was recorded (Last et al., 2011).  The lower mortality when buried in coarse sands may be related to the greater number of individuals who were able to emerge in these conditions and emergence was to be significant for survival. 

It is unclear whether the same results would be recorded when mussels are joined by byssal threads or whether this would have an impact on survival (Last et al., 2011), although Daly & Mathieson (1977) recorded loose attachments between juvenile mussels during a burial event and some of these were able to surface.  It was not clear whether the same ability would be shown by adult mussels in a more densely packed bed.

Sensitivity assessment. Overburden by 5 cm of fine material (see benchmark) in a single incident is unlikely to result in significant mortality in blue mussel beds before sediments are removed by current and wave action.  However, the inability of Mytilus edulis to emerge from sediment deeper than 2 cm (Last et al., 2011, Essink, 1999, Daly & Matthieson, 1977) and the increased mussel mortality with depth and reduced particle size observed by Last et al. (2011) suggest that there may be some mortality and resistance is assessed as 'Medium'. Survival will be higher in winter months when temperatures are lower and physiological demands are decreased.  However, mortality will depend on the duration of smothering. Mortality is likely to be more significant in wave sheltered areas, devoid of tidal streams, where the smothering sediment remains for prolonged periods.  Mortality will be more limited, and possibly avoided, where the smothering sediment is removed due to wave action or tidal streams, depending on how long the sediment remains over the individual mussels.  Resilience is assessed as 'Medium' and sensitivity is, therefore 'Medium'. 

Smothering and siltation rate changes (heavy)

Benchmark. ‘Heavy’ deposition of up to 30 cm of fine material added to the seabed in a single discrete event. Further detail

Evidence

The main human activity that increases sedimentation is dredging and the dumping of dredged sediments in estuarine and coastal waters.  Aggregate dredging and fishing gear can cause localised sedimentation and smothering.  However, changes in water flow can cause localised smothering within mussel beds (Widdows et al., 2002), and storms may move large volumes of sediment and smother entire mussel beds (Dare, 1976). Mytilus edulis occurs in areas of high suspended particulate matter (SPM) and therefore a level of siltation is expected from the settling of SPM.  In addition, the high rate of faecal and pseudofaecal matter production by the mussels naturally results in siltation of the seabed, often resulting in the formation of large mounds beneath the mussel bed.  For example, at Morecambe Bay, an accumulation of mussel-mud (faeces, pseudofaeces and washed sand) of 0.4-0.5 m between May 1968 and September 1971 resulted in the mortality of young mussels (Daly & Mathieson, 1977).  In order to survive the mussels needed to keep moving upwards to stay on the surface.  Many individuals did not make it to the surface and were smothered by the accumulation of mussel-mud (Daly & Mathieson, 1977), so that whilst Mytilus edulis does have the capacity to vertically migrate through sediment some individuals will not survive.  

Sand burial has been shown to determine the lower limit of Mytilus edulis beds (Daly & Mathieson, 1977a).  Burial of Mytilus edulis beds by large-scale movements of sand, and resultant mortalities have been reported from Morecambe Bay, the Cumbrian coast and Solway Firth (Holt et al., 1998).  Essink (1999) recorded fatal burial depths of 1-2 cm for Mytilus edulis and suggested that they had a low tolerance of sedimentation based on investigations by R.Bijkerk (cited by Essink, 1999).  Essink (1999) suggested that deposition of sediment (mud or sand) on shallow mussel beds should be avoided.  However, Widdows et al. (2002) noted that mussels buried by 6 cm of sandy sediment (caused by resuspension of sediment due to turbulent flow across the bed) were able to move to the surface within one day.  Conversely, Condie (2009) (cited by Last et al., 2011) reported that Mytilus edulis was tolerant of repeated burial events. 

Last et al., (2011) carried out burial experiments on Mytilus edulis in pVORTs.  They used a range of burial depths and sediment fractions and temperatures.  It was found that individual mussels were able to survive burial in depths of 2, 5 and 7 cm for over 32 days although the deeper and longer the mussels were buried the higher the mortality.  Only 16% of buried mussels died after 16 days compared to almost 50% mortality at 32 days.  Mortality also increased sharply with a decrease in particle size and with increases in temperature from 8.0 and 14.5 to 20 °C.  The ability of a proportion of individuals to emerge from burial was again demonstrated with approximately one quarter of the individuals buried at 2 cm resurfacing.  However, at depths of 5 cm and 7 cm no emergence was recorded (Last et al., 2011).  The lower mortality when buried in coarse sands may be related to the greater number of individuals who were able to emerge in these conditions and emergence was to be significant for survival. 

It is unclear whether the same results would be recorded when mussels are joined by byssal threads or whether this would have an impact on survival (Last et al., 2011), although Daly & Mathieson (1977) recorded loose attachments between juvenile mussels during a burial event and some of these were able to surface.  It was not clear whether the same ability would be shown by adult mussels in a more densely packed bed.

Sensitivity assessment. Overburden by 30 cm of fine material (see benchmark) in a single incident could result in significant mortality in blue mussel beds due to the limited ability of Mytilus edulis to emerge from sediment deeper than 2 cm (Last et al., 2011, Essink, 1999, Daly & Matthieson, 1977) and the increased mussel mortality with depth and reduced particle size observed by Last et al. (2011).  Survival will be higher in winter months when temperatures are lower and physiological demands are decreased. However, mortality will depend on the duration of smothering. Mortality will be limited, and possibly avoided, where the smothering sediment is removed due to wave action or tidal streams, depending on how long the sediment remains over the individual mussels. However, mortality is likely to be significant in wave sheltered areas, devoid of tidal streams, where the smothering sediment remains for prolonged periods (e.g. more than 16 days).  Therefore, resistance has been assessed as ‘Low’ (significant mortality, loss of 25-75 % of population abundance, or extent) for the littoral mussel sediment biotopes (LS.LBR.LMus and LS.LBR.LMus.Myt).   Hence, resilience is assessed as ‘Medium’, so that sensitivity is 'Medium'.  

Litter

Benchmark. The introduction of man-made objects able to cause physical harm (surface, water column, seafloor or strandline). Further detail

Evidence

Not assessed.

Electromagnetic changes

Benchmark. A local electric field of 1 V/m or a local magnetic field of 10 µT. Further detail

Evidence

No evidence.

Underwater noise changes

Benchmark. MSFD indicator levels (SEL or peak SPL) exceeded for 20% of days in a calendar year. Further detail

Evidence

Not relevant.

Introduction of light or shading

Benchmark. A change in incident light via anthropogenic means. Further detail

Evidence

No evidence.

Barrier to species movement

Benchmark. A permanent or temporary barrier to species movement over ≥50% of water body width or a 10% change in tidal excursion. Further detail

Evidence

No direct evidence was found to assess this pressure. As the larvae of Mytilus edulis are planktonic and are transported by water movements, barriers that reduce the degree of tidal excursion may alter the supply of Mytilus edulis to suitable habitats from source populations. However, the presence of barriers may enhance local population supply by preventing the loss of larvae from enclosed habitats.  This species is therefore potentially sensitive to barriers that restrict water movements, whether this will lead to beneficial or negative effects will depend on whether enclosed populations are sources of larvae or are ‘sink’ populations that depend on outside supply of larvae to sustain the local population.

Sensitivity assessment. As this habitat is potentially sensitive to changes in tidal excursion and exchange, resistance is assessed as ‘Medium’ and resilience as ‘High’, sensitivity is, therefore ‘Low’.

Death or injury by collision

Benchmark. Injury or mortality from collisions of biota with both static or moving structures due to 0.1% of tidal volume on an average tide, passing through an artificial structure. Further detail

Evidence

Not relevant to benthic habitats.  NB. Collision by grounding vessels is addressed under ‘surface abrasion’.

Visual disturbance

Benchmark. The daily duration of transient visual cues exceeds 10% of the period of site occupancy by the feature. Further detail

Evidence

Not relevant.

Genetic modification & translocation of indigenous species

Benchmark. Translocation of indigenous species or the introduction of genetically modified or genetically different populations of indigenous species that may result in changes in the genetic structure of local populations, hybridization, or change in community structure. Further detail

Evidence

Commercial cultivation of Mytilus edulis involves the collection of juvenile mussel ‘seed’ or spat (newly settled juveniles ca 1-2 cm in length) from wild populations, with subsequent transportation around the UK for re-laying in suitable habitats. As the seed is harvested from wild populations from various locations the gene pool will not necessarily be decreased by translocations.  Movement of mussel seed has the potential to transport pathogens and non-native species (see relevant pressure sections). This pressure assessment is based on Mainwaring et al. (2014) and considers the potential impacts on natural mussel beds of genetic flow between translocated stocks and wild mussel beds. 

Commercial cultivation of Mytilus edulis involves the collection of juvenile mussel ‘seed’ or spat (newly settled juveniles ca 1-2 cm in length) from wild populations, with subsequent transportation around the UK for re-laying in suitable habitats. As the seed is harvested from wild populations from various locations the gene pool will not necessarily be decreased by translocations.  Movement of mussel seed has the potential to transport pathogens and non-native species (see relevant pressure sections). This pressure assessment is based on Mainwaring et al. (2014) and considers the potential impacts on natural mussel beds of genetic flow between translocated stocks and wild mussel beds. 

Two species of Mytilus occur in the UK, Mytilus edulis and Mytilus galloprovincialis.  Mytilus edulis appears to maintain genetic homogeneity throughout its range whereas Mytilus galloprovincialis can be genetically subdivided into a Mediterranean group and an Atlantic group (Beaumont et al. 2007).  Mytilus edulis and Mytilus galloprovincialis have the ability to hybridize in areas where their distribution overlaps e.g. around the Atlantic and European coast (Gardner, 1996; Daguin et al., 2001; Bierne et al., 2002; Beaumont et al., 2004).  In the UK overlaps occur on the North East coast, North East Scotland, South West England and in the North, West and South of Ireland (Beaumont et al., 2007).  It is difficult to distinguish Mytilus edulis, Mytilus galloprovincialis or hybrids based on shell shape because of the extreme plasticity of shape exhibited by mussels under environmental variation, and a genetic test is required (Beaumont et al., 2007).  There is some discussion questioning the distinction between the two species as the hybrids are fertile (Beaumont et al., 2007).  Hybrids reproduce and spawn at a similar time to both Mytilus edulis and Mytilus galloprovincialis which supports genetic flow between the taxa (Doherty et al., 2009).

There is some evidence that hybrid larvae have a faster growth rate to metamorphosis than pure individuals which may leave pure individuals more vulnerable to predation (Beaumont et al., 1993).  As the physiology of both the hybrid and pure Mytilus edulis is so similar there is likely to be little impact on the tolerance of the bed to either pressure nor a change in the associated fauna. A review by Svåsand et al. (2007) concluded that there was a lack of evidence distinguishing between different populations to accurately assess the impacts of hybridization and in particular how the gene flow may be affected by aquaculture.  Therefore, it cannot be confirmed whether farming will have an impact on the genetics of this species beyond a potential for increased hybridization.

Sensitivity assessment. No direct evidence was found regarding the potential for negative impacts of translocated mussel seed on adjacent natural beds.  While it is possible that translocation of mussel seed could lead to genetic flow between cultivated beds and local wild populations, there is currently no evidence to assess the impact (Svåsand et al., 2007).  Hybrid beds perform the same ecological functions as Mytilus edulis so that any impact relates to genetic integrity of a bed alone.  This impact is considered to apply to all mussel bed biotopes equally, as the main habitat-forming species Mytilus edulis is translocated.  Also, given the uncertainty in identification of the species, habitats or biotopes described as dominated by Mytilus edulis may well be dominated by Mytilus galloprovincialis, their hybrids or a mosaic of the three. Presently, there is 'No evidence' of impact due to genetic modification and translocation; therefore ‘No evidence’ is reported.  The range of Mytilus galloprovincialis is thought to be extending northwards (Beaumont et al., 2007) and this assessment may require updating in the future.  

Introduction or spread of invasive non-indigenous species

Benchmark. The introduction of one or more invasive non-indigenous species (INIS). Further detail

Evidence

Sewell et al. (2008) reviewed the evidence for invasive non-indigenous species with the potential to be introduced to and impact mussel beds.  These included Botrylloides violaceus, Corella eumyota, Crepidula fornicata, Didemnum vexillum, Rapana venosa, Magallana gigas and Aulocomya ater. 

The Pacific oyster, Magallana (syn. Crassostrea) gigas, is native to warm temperate regions from the northwest Pacific to Japan and northeast Asia, including Cape Mariya (Russia) to Hong Kong (China) (Carrasco & Baron, 2010; GBNNSS, 2011, 2012). It is a fast-growing and tolerant species that has become a successful invader in the coastal waters of all continents, aside from Antarctica (Wrange et al., 2010; Carrasco & Baron, 2010; Padilla, 2010). Magallana gigas is recognised as a beneficial and important species in aquaculture worldwide (Padilla, 2010). It was initially introduced for aquaculture in Europe and the UK in the 1960s due to a decline in the Portuguese oyster (Crassostrea angulata) and the European flat oyster (Ostrea edulis) (Spencer et al., 1994; GBNNSS, 2011, 2012; Humphreys et al., 2014 cited in Alves et al., 2021; Hansen et al., 2023).

Since introduction, the species has invaded and established self-sustaining natural populations throughout Europe from the North Sea, Wadden Sea and Scandinavian coastlines to the Atlantic coastlines of Spain and Portugal, as well as the Mediterranean and Adriatic Sea (Wrange et al., 2010; GBNNSS, 2011, 2012; Ezgeta-Balic et al., 2019; Spagnolo et al., 2019; Bergstrom et al., 2021; Hansen et al., 2023). In the UK, the species predominantly occurs around the southern and western coastlines (OBIS, 2024; NBN, 2024). Shipping activity has also been associated with the introduction of Magallana gigas in the northeastern Adriatic Sea, where it was not introduced for aquaculture (Ezgeta-Balic et al., 2019) and possibly in southwest England from France possibly via fouling on ships (GBNNSS, 2011, 2012; Padilla, 2010; Ezgeta-Balic et al., 2019).

Magallana gigas has a high fecundity, a long-lived pelagic larval phase (2 to 4 weeks) and can produce up to 200 million eggs during spawning (Herbert et al., 2012, 2016; Alves et al., 2021; Wood et al., 2021; Hansen et al., 2023). Hence, as a broadcast spawner, it has a high dispersal potential of more than 1000 km (Padilla, 2010; Wood et al., 2021). Larval mortality can be as large as 99%, as larvae are sensitive to environmental conditions (Alves et al., 2021), but adults are long-lived so populations can survive with infrequent recruitment (Padilla, 2010). Larval dispersal and mass spawning events have facilitated the settlement and establishment of Pacific oysters, as seen in the Oosterschelde estuary, Netherlands (Hansen et al., 2023). It has been suggested that the spread of the Pacific oyster in Scandinavia is due to northward larval drift on tidal and wind-driven currents (Hansen et al., 2023). Wood et al. (2021) suggested that larval dispersal of the Pacific oyster from populations within and outside the UK was possible via unaided (passive) transport by currents, but that aquaculture and offshore structures (e.g. windfarms) increased the risk of the invasive species spreading and the geographical extent of spread.

Pacific oyster reefs, in the Wadden Sea and Brittany, on littoral muddy and sandy habitats formed predominantly at lower tidal levels from Mean Low Water levels to the shallow subtidal (Herbert et al., 2012, 2016). Pacific oyster spatfall was recorded in the estuarine intertidal zone on areas with hard substrata of stone and shell, particularly between the low water of spring tides and high water of neap tides, such as in the Menai Strait (Spencer et al., 1994). In Lim Bay, Adriatic Sea, Magallana gigas is only found in the intertidal and on the sublittoral edge (at a depth of 1 m) and not at 3 m or 6 m depth (Stagličić et al., 2020; Tillin et al., 2020). It coexists here with Ostrea edulis which is abundant in the subtidal (Stagličić et al., 2020). Bergstrom et al. (2021) found that depth was one of the most important predictors of the occurrence of Magallana gigas in the Skagerrak and suggested the optimal depth of the species was 0.5 m in the shallow subtidal, although it occurred down to 5 m. The majority of the evidence indicates that infralittoral rock and other habitats that occur at depths more than 10 m are unlikely to be suitable for Magallana gigas because it is considered an intertidal and shallow subtidal species rarely recorded below extreme low water (Herbert et al., 2012, 2016; Tillin et al., 2020). However, in suitable situations (e.g. Oosterschelde) it may form beds down to 42 m.

In the Wadden Sea and the North Sea, Magallana gigas overgrows mussel beds in the intertidal zone, on sedimentary and rocky habitats of low or high energy (Diederich, 2005, 2006; Nehls et al., 2006; Kochmann et al., 2008; Wrange et al., 2010; Padilla, 2010; GBNNSS, 2011, 2012; Kochmann, 2012; Kochmann et al., 2013; Herbert et al., 2016; Tillin et al., 2020). The Pacific oyster can out-compete Mytilus edulis, particularly for food and space, as the faster growth rates of the oyster make it more competitive when food or space is limiting (Nehls et al., 2006; Padilla, 2010; Tillin et al., 2020; Joyce et al., 2021). For example, in Sylt, Wadden Sea, mudflats and mussel beds have now been changed into Magallana gigas reefs (Tillin et al., 2020). In the northern Wadden Sea, this change is considered permanent (Tillin et al., 2020).

Diederich (2005, 2006) examined the settlement, recruitment, and growth of Magallana gigas and Mytilus edulis in the northern Wadden Sea. Magallana gigas recruitment success was dependent on temperature, and in the northern Wadden Sea, only occurred in six of the 18 years since Magallana gigas was first introduced. Survival of juveniles is higher in mild than in cold winters. Also, the survival of both juveniles and adults on mussel beds is higher than that of the mussels themselves. However, recruitment of Magallana gigas was significantly higher in the intertidal than the shallow subtidal, although the survival of adult oysters or mussels in the subtidal is limited by predation. Deiderich (2005) concluded that hot summers could favour Magallana gigas reproduction while cold winters could lead to high mussel recruitment the following summer. Diederich (2005, 2006) noted that the high survival rate of Magallana gigas adults and juveniles in the intertidal was likely to compensate for years of poor recruitment. Magallana gigas also prefers to settle on conspecifics, so that it can build massive oyster reefs, which themselves are more resistant to storms or ice scour than the mussel beds they replace, as oysters are cemented together, rather than dependent on byssus threads. Magallana gigas also grows faster than Mytilus edulis in the intertidal and reaches ca 2-3 times the length of mussels within one year. In addition, growth rates in Magallana gigas were independent of the tidal level (emergence regime, substratum, Fucus cover and barnacle epifauna (growing on both mussels and oysters), while the growth rate of Mytilus edulis was decreased by these factors. The faster growth rate could make Magallana gigas more competitive than Mytilus edulis where space or food is limiting. Diederich (2006) concluded that the massive increase in Magallana gigas in the northern Wadden Sea was caused by high recruitment success, itself due to anomalously warm summer temperatures, the preference for settlement on conspecifics (and hence reef formation), and high survival rates of juveniles. As oyster reefs form on former mussel beds, the available habitat for Mytilus edulis could be restricted (Diederich, 2006). In addition, in the northern German Wadden Sea, the decrease in blue mussel beds and increase in Pacific oysters was linked to climatic conditions rather than caused by the invasion of the Pacific oyster (Nehls et al., 2006).

Kent and Essex Inshore Fisheries and Conservation Authority (IFCA) (cited in Herbert et al., 2012) reported that Magallana gigas had developed a significant stock on mussel beds on the Southend foreshore and that, by 2012, there were few mussels left in the affected area but made no conclusions as to the reason for the decline in mussels. Herbert et al. (2016) reported that many Mytilus edulis beds have changed into mixed reefs dominated by 95% Magallana gigas in the Wadden Sea.

Despite concerns that the Pacific oyster can out-compete the Mytilus edulis, research indicates that mixed reefs can shift densities of resident species without suppressing native mussels and the two species can coexist as mixed ‘oyssel’ beds (Reise et al., 2017; Cornelius & Buschbaum, 2020; Joyce et al., 2021). The invasion of Magallana gigas may alter the structure and function of intertidal reefs in the short-term but can sometimes create a multi-layered structure of a mixture of oysters and blue mussels in the long-term that is more resilient and accumulates a higher biodiversity of flora and fauna and supports the densities of other native species such as Littorina littorea (Andriana et al., 2020; Cornelius & Buschbaum, 2020). Reise et al. (2017) noted that in the initial stage of colonization, oysters used mussels for settlement and smothered the bed. Ten years later, the oyster bed became the preferred substratum for settlement and, after 20 years, mussels were no longer the preferred substratum for oyster larvae and were able to use the oyster bed to shelter from predation and parasites (Reise et al., 2017). However, on the remaining hummocks of mussel mud, mussels dominate the top of the hummock and oysters on the sides (Reise et al., 2017). Native and invasive oysters are known to provide a refuge from predators within the biogenic reef they create (Troost, 2010; Goedknegt et al., 2020). The blue mussel Mytilus edulis can make use of shelter provided by the Pacific oysters to escape predators by migrating to the bottom of the Pacific oyster reef reducing mussel predation by crabs and birds (Goedknegt et al., 2020). Therefore, the presence of Magallana gigas in mussel beds can adjust the mussel predator avoidance. Mixed oyster-mussel beds (‘oyssel’ beds) were reported to exhibit increased species richness, abundance, biomass, and number of deposit feeders compared to mussel beds in the German Wadden Sea (Markert et al., 2010; Herbert et al., 2016; Cornelius & Buschbaum, 2020).

The global spread of the Pacific oyster has facilitated the introduction of macrospecies and parasites associated with oysters, including harmful algae and disease agents (Padilla, 2010). It is recognised that copepod parasites of Magallana gigas, Mytilicola orientalis and Myicola ostreae were introduced with imports of the Pacific oyster from France to Ireland (Tillin et al., 2020). Mytilicola orientalis was introduced into the Wadden Sea by Magallana gigas and infected blue mussels (Goedknegt et al., 2020). Predator avoidance by blue mussels in biogenic oyster reefs can indirectly affect parasite-host interactions. For example, in the Wadden Sea, one mixed mussel and oyster reef had a significantly higher abundance of parasitic Mytilicola spp. in mussels at the top of the reef compared to those at the bottom (Goedknegt et al., 2020). In contrast, with increasing oyster density, an increase in the presence of the trematode Renicola roscovita was seen in mussels (Goedknegt et al., 2019). Magallana gigas is also the predominant host of the shell-boring parasites Polydora ciliata and Polydora websteri in the Wadden Sea, with relatively higher densities of Polydora ciliata found in the Pacific oyster compared to the blue mussels (Waser et al., 2021).

The American slipper limpet Crepidula fornicata was introduced to the UK and Europe in the 1870s from the Atlantic coasts of North America with imports of the eastern oyster Crassostrea virginica. It was recorded in Liverpool in 1870 and the Essex coast in 1887-1890. It has spread through expansion and introductions along the full extent of the English Channel and into the European mainland (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 1999, 2018; Hinz et al., 2011; Helmer et al., 2019; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015). Crepidula fornicata is reported to settle and establish amongst mussel beds (Blanchard, 1997; Thieltges, 2005; Rayment, 2007).  If Crepidula fornicata becomes established in a bed it is likely to alter the bed structure particularly if it is on coarse sand or hard substrata. Crepidula fornicata has high fecundity and can disperse its larvae over large areas making mussel beds highly vulnerable if Crepidula fornicata is introduced even large distances away.  The larvae of Crepidula fornicata can survive transport in ballast water for a number of days allowing it to travel large distances before needing to settle in the areas where the ballast water is released (Blanchard, 1997).  Crepidula can colonize a wide range of substrata. It prefers muddy gravelly, shell-rich, substrata that include gravel, or shells of other Crepidula, or other species e.g., oysters, and mussels. It is highly gregarious and seeks out adult shells for settlement, forming characteristic ‘stacks’ of adults. But it also recorded from rock, artificial substrata, and Sabellaria alveolata reefs (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Hinz et al., 2011; Helmer et al., 2019; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015; Tillin et al., 2020). 

Thieltges et al. (2003) reported that Crepidula fornicata was abundant on mussel beds in the intertidal to subtidal transition zone, in the northern Wadden Sea in the year 2000. Crepidula had increased in abundance since 1948 and had expanded its range from the extinct oyster beds to mussel beds where live mussels were its main substratum. Thieltges et al. (2003) also noted that storm events removed some clumps of mussels and presumably Crepidula onto tidal flats where they disappeared, which caused their abundance to fluctuate. Thieltges et al. (2003) noted that Crepidula abundance at the intertidal to the subtidal transition zone (ca 21 /m²) was significantly higher than in the upper, mid, and lower intertidal (ca <3 /m²). Thieltges (2005) reported a 28-30% mortality of Mytilus edulis when Crepidula fornicata was introduced to the beds in experimental studies. He also found that mussel shell growth was reduced by 3 to 5 times in comparison to unfouled mussels and that extra energy was probably expended on byssus production.  The most significant cause of mortality was increased drag on mussels due to the growth of stacks of Crepidula fornicata on the shells of the mussels, rather than competition for food. He concluded that Crepidula fornicata is potentially an important mortality factor for Mytilus edulis (Thieltges, 2005).  Thieltges (2005) also observed mussel beds in the shallow subtidal infested with high abundances of Crepidula fornicata with almost no living mussels, along the shore of the List tidal basin, northern Wadden Sea.  

The density of Crepidula populations in northern Europe (Germany, Denmark, and Norway) is significantly lower (ca <100 /m²) than in southern waters. Thieltges et al. (2004) reported that the population of Crepidula was affected strongly by cold winters in the Wadden Sea. The winters of 2001 and 2003 resulted in ca 56-64% mortality of intertidal Crepidula and up to 97% on one mussel bed, compared to only 11-14% in southern areas without frost. Crepidula almost vanished from the Wadden Sea after the 1978/79 winter and took ten years to recover due to moderate winters which regularly affected the population. Similarly, 25% mortality was observed in Crepidula populations on the south coast of the UK after the extreme 1962/63 winter (Crisp, 1964, Bohn et al., 2012). Thieltges et al. (2003) suggested that global warming may allow Crepidula populations to become more abundant in northern Europe. Valdizan et al. (2011) noted higher water temperatures between 2000 to 2001 and 2006 to 2007 together with elevated chlorophyll-a corresponded to an increase in gametogenesis and the duration of broods in the Crepidula population in Bournerf Bay, France. They suggested that rising temperatures in northern Europe could increase its reproductive success due to favourable breeding temperatures and increased phytoplankton (Valdizan et al., 2011). Nehls et al. (2006) noted that the decline in mussel (Mytilus edulis) beds in the Wadden Sea was due to mild winters that favoured non-native oysters (Magellana gigas) and slipper limpets, which co-existed with the mussels.

Bohn et al. (2013a) reported that mussel shells provided a more suitable settlement substratum for Crepidula larvae than bare panels in larval settlement experiments. However, the presence of live Mytilus edulis did not increase colonization of the site by Crepidula in the Milford Harbour Waterway, e.g., no Crepidula were found on mussels at a site with 23% cover of mussels (Bohn et al., 2015). Bohn et al. (2015) suggested that its prevalence on mussels in the Wadden Sea was due to a lack of alternative substratum, together with the cold weather mortalities. 

Crepidula fornicata is likely to alter water flow over mussel beds.  They form stacks of individuals that change water flow across the sediment surface.  When these stacks occur on the shells of Mytilus edulis they increase the drag on the mussel, increase the demands on the mussel’s energy reserves for attachment (e.g. byssus formation) and, hence, affect fecundity and survival (Thieltges, 2005; Sewell et al., 2008).  The increased drag may also result in clumps of mussels being removed by water flow (Thieltges, 2005).  Competition for suspended organic matter and space is also increased.  Space for the settlement of macrobenthic organisms (Blanchard, 1997) including mussels is particularly reduced.  In addition to the reduced space for settlement, larvae of macrobenthic organisms are consumed by the slipper limpet and may affect recruitment to an area. 

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009). Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024). In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; Michin & Nunn, 2013; Bishop et al., 2015; Mckenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

Although a widespread invader, Didemnum vexillum has a limited ability for natural dispersal since the pelagic larvae remain in the water column for a short time (up to 36 hours). Therefore, it has a short dispersal phase that can allow the species to build localized populations (Herborg et al., 2009; Vercaemer et al., 2015; Holt, 2024). However, Bullard et al. (2007) suggested that Didemnum vexillum can form new colonies asexually by fragmentation. Colonies can produce long tendrils from an encrusting colony, which can fragment, disperse and settle, attaching to suitable hard substrata elsewhere (Bullard et al., 2007; Lambert, 2009; Stefaniak & Whitlatch, 2014). A fragmented colony can spread naturally for up to three weeks transported by ocean currents, attached to floating seaweed, seagrass or other floating biota, or as free-floating spherical colonies (Bullard et al., 2007; Lengyel et al., 2009; Stefaniak & Whitlatch, 2014; Holt, 2024). Fragments can reattach to suitable substrata within six hours of contact. Fragments have the potential to disperse around 20 km before reattachment (Lengyel et al., 2009). Valentine et al. (2007a) reported that colonies of Didemnum vexillum enlarged by 6 to 11 times by asexual budding after 15 days and enlarged 11 to 19 times after 30 days. Valentine et al. (2007a) concluded fragments could successfully grow, survive, and help to spread Didemnum vexillum.

While natural fragmentation of tendrils is thought to allow Didemnum vexillum to invade longer distances and increase its dispersal potential, Stefaniak & Whitlatch (2014) found that only a one tendril out of 80 reattached to the flat, bare substrata used in their study, because tendrils required an extensive (at least eight hour) period of contact to reattach. Stefaniak & Whitlatch (2014) suggested that once fragmented from a colony, the success of tendril reattachment was limited, and reattachment was not a major contributor to the invasive success of Didemnum vexillum. However, Stefaniak & Whitlatch (2014) also found that larvae-packed tendril fragments may increase natural dispersal distance, reproduction, and invasive success of Didemnum vexillum, and increase the distance larvae can travel. Not all colonies produce tendrils at all locations.

Human-meditated transport via aquaculture facilities, boat hulls, commercial fishing vessels, and ballast water is probably the most important vector that has aided the long-distance dispersal of Didemnum vexillum and explains its prevalence in harbours and marinas (Bullard et al., 2007; Dijkstra et al., 2007; Griffith et al., 2009; Herborg et al., 2009). Fragmentation of colonies during transport or human disturbance (such as trawling or dredging) could indirectly disperse the species and enable it to find suitable conditions for establishment (Herborg et al., 2009). For example, in oyster farms in British Columbia, large fragments of Didemnum sp. come off oyster strings when they are pulled out of water and other fragments can be pulled off oysters and mussels and thrown back into the water, which is likely to aid dispersal of the invasive species (Bullard et al., 2007). Dijkstra et al. (2007) hypothesised that Didemnum sp. was introduced to the Gulf of Maine with oyster aquaculture in the Damariscotta River and transported via Pacific oysters.

Didemnum vexillum was likely introduced into the UK from northern Europe or Ireland via poorly maintained or not antifouled vessels, movement of contaminated shellfish stock and aquaculture equipment, or via marine industries such as oil, gas, renewables, and dredging (Holt, 2024). Recent evidence from genetic material suggests that human-mediated dispersal, between marinas and shellfish culture sites, is the most likely pathway for connectivity of Didemnum vexillum populations throughout Ireland and Britain (Prentice et al., 2021; Holt, 2024). Didemnum vexillum can disperse away from artificial substrata, invading and colonizing natural substrata in surrounding areas (Tillin et al., 2020). Holt (2024) noted that Didemnum vexillum had not spread as far as feared in the UK since it was first recorded. The current evidence of Didemnum vexillum’s ability to spread on natural habitats in this area is sparse and often conflicting, complicated by genetics and its apparent variable habitat preferences and tolerances and its variable ability to adapt to ‘new’ conditions (Holt 2024).

Didemnum vexillum has a seasonal growth cycle that is influenced by temperature (Valentine et al., 2007a). In warmer months (June and July) colonies may be large and well-developed encrusting mats. Populations experience more rapid growth from July to September sometimes continuing into December. Colonies begin to decline in health and ‘die-off’ when temperatures drop below 5°C during winter months from around October to April (Gittenberger, 2007; Valentine et al., 2007a; Herborg et al., 2009). Cold water months cause colonies to regress and reduce in size, yet they often regenerate as temperatures warm (Griffith et al., 2009; Kleeman, 2009, Mercer et al., 2009), although some populations may not survive winter at all (Dijkstra et al., 2007). The early growth phase, from May to July, is initiated by smaller colonies developing from remnants of colonies that survived the cold water (Valentine et al., 2007a). The seasonal growth cycle is also likely influenced by location. For example, the Didemnum sp. growth cycle for colonies in Sandwich tide pool (temperature range from -1 °C to 24 °C, with daily fluctuations), probably does not occur in deep offshore subtidal habitats in Georges Bank (annual temperature range from 4 °C to 15°C, and daily fluctuations are minimal) (Valentine et al., 2007a).  Larval release and recruitment typically occur between 14 to 20°C and slow or cease below 9 to 11°C as summer ends (Griffith et al., 2009; Mckenzie et al., 2017). In New Zealand, recruitment occurs from November to July, where the highest average temperatures were recorded in February (18 to 22°C) and the lowest average temperatures were recorded in July (9 to 10°C) (Fletcher et al., 2013a). In this New Zealand study, higher water temperatures were associated with a higher level of recruitment (Fletcher et al., 2013a).

Didemnum vexillum requires suitable hard substrata for successful settlement and the establishment of colonies. It can grow quickly and establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems because of its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock gravel, pebble, cobble, or boulders or artificial substrata such as a variety of maritime structures such as pontoons, docks, wood and metal pilings, chains, ropes and moorings, plastic and ship hulls and at aquaculture facilities (Valentine et al., 2007 a&b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012; Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent, and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020).

There are few observations of Didemnum vexillum on soft bottom habitats as evidence suggests it is unable to establish or grow easily on mud, mobile sand or other unstable substrata, and it is vulnerable to smothering by fine sediment (Bullard et al., 2007; Valentine et al., 2007a; Griffith et al., 2009). The species is usually found established in areas where the colony is protected from sedimentation and wave action (Valentine et al., 2007b; Mckenzie et al., 2017; Tillin et al., 2020). For example, at Georges Bank, USA the Didemnum vexillum mats were limited to gravelly areas and unable to colonize the sand ridges that bounded the site, which have a mobile surface that is moved daily by the strong tidal currents (Valentine et al., 2007b). In addition, evidence found the species can also not survive being buried or smothered by coarse or fine grained sediment. Furthermore, in Holyhead marina, Didemnum vexillum colonies were contained in the harbour and established on artificial pontoons, and they were not present on the natural seabed under the pontoon, which is composed of silty mud or on deeper sections of mooring chains that are immersed in mud at low spring tides (Griffith et al., 2009).

However, some studies on Georges Bank, USA and Sandwich, Massachusetts observed colonies were able to survive partial covering by sand (Bullard et al., 2007; Valentine et al., 2007a). Gittenberger et al. (2015) reported that Didemnum vexillum was able to overgrow sandy bottom (cited Gittenberger, 2007). In northern Kent, Didemnum vexillum has been recorded covering London clay boulders on Whitestable Flats, West Beach, north Kent, covering tabulate sandstone boulders (0.5 to 2 m across) on the mid-shore and colonizing sediment mounds on the low shore characterized by larger areas of sand, mud and low-lying sediment at Reculver and Bishopstone, north Kent (Hitchin, 2012). It was also recorded from muddy substrata at that site. Hitchin (2012) noted that the site was exposed to enough waves and currents to cause sedimentation. However, Didemnum vexillum grew hanging from on the underside of sandstone boulders nestled on sediment, on consolidated sediment mounds and firm clays, hence burial may prevent colonization and its survival rather than sedimentation alone.

In contrast, Didemnum vexillum’s preference for sheltered conditions, established colonies observed in Georges Bank and Long Island Sound were exposed to moderately strong tidal currents (1 to 2 knots; ca 0.5 to 1 m/s recorded at both sites) that may mobilise sediment (Valentine et al., 2007b; Mercer et al., 2009; Tillin et al., 2020). However, Valentine et al. (2007b) describe the substratum as immobile, presumably consolidated, gravel, cobbles, and pebbles. Kleeman (2009), stated that the presence of a consistent mild wave action or ‘swash zone’ appears to favour Didemnum sp. establishment in the intertidal. Although some evidence suggests that waves and currents can facilitate the fragmentation and spread of Didemnum vexillum (Mckenzie et al., 2017), the tidal current velocities at some sites where Didemnum vexillum has been reported (for example, New England, where current velocities reach up to around 3 m/s) is lower than the current velocity required for the dislodgement of Didemnum vexillum fragments (around 7.6 m/s) (Reinhardt et al., 2012). This suggests that not all tidal currents are likely to dislodge Didemnum vexillum fragments. When on boat hulls the species can experience higher current velocities which is enough to cause dislodgement (Reinhardt et al., 2012).  

Didemnum vexillum can overgrow bivalve species, such as oysters, scallops, and mussels, as the hard shells can provide suitable hard substrata for settlement. It has been described as a ‘shellfish pest’ by the aquaculture industry because it is likely to completely encapsulate bivalves and smother them resulting in death or partially encapsulate and partially smother them resulting in reduced bivalve growth (Auker, 2010; Bullard et al., 2007; Coutts & Forrest, 2007, Valentine et al., 2007a; Carman et al., 2009; Kleeman, 2009; Fletcher et al., 2013b; Tillin et al., 2020;). Didemnum vexillum has been recorded overgrowing mussels in Strangford Lough, Northern Ireland (Minchin & Nunn, 2013) and recorded forming large mats over Blue Mussel beds in the Gulf of Maine, completely covering individuals (Auker et al., 2014). Didemnum vexillum fouling on aquaculture equipment and bivalve species causes great economic impacts, as Didemnum vexillum removal methods are expensive, labour-intensive, and not always effective (Coutts & Forrest, 2007; Carman et al., 2009; Kleeman, 2009; Fletcher et al., 2013b; Tillin et al., 2020; Holt, 2024). The fouling on aquaculture nets and bags can restrict water flow and food availability for shellfish and smothering on mussel farms may result in crop losses (Coutts & Forrest, 2007; Carver et al., 2003 cited by Carman et al., 2009; Fletcher et al., 2013b; Holt, 2024). Effects on mussels are likely to become more prominent as Didemnum vexillum becomes more abundant (Auker, 2010).

The epibiotic relationship between Didemnum vexillum and Mytilus edulis negatively impacts mussel growth (Auker, 2010). Clean control mussels with no Didemnum vexillum overgrowth had thicker shells, a significantly thicker lip, and a greater tissue index, compared to mussels overgrown by Didemnum vexillum (Auker, 2010). The clean mussels’ average length ranged from 3.2 cm to 5.37 cm, and had significantly greater shell lengths than overgrown mussels, which had an average length from 3.4 cm to 4.86 cm (Auker, 2010). Mortality of both control and overgrown mussels was relatively low over the one-year study period, but higher mortality was seen in overgrown mussels (6.7% died) compared to the clean control mussels (1.1% died) (Auker, 2010). Food is an important factor contributing to the decrease in mussel growth (Auker, 2010). Auker (2010) also found that Didemnum vexillum affected reproduction and recruitment of Mytilus edulis as the invasive species grew over gamete release point (siphons) or inhibited settlement of recruits, but this varied seasonally.

In contrast, the overgrowth of mussels by Didemnum vexillum has reduced the predation risk on mussels (Auker, 2010; Auker et al., 2014, Lyu et al., 2020). The Didemnum vexillum mats act as refuges for blue mussels (Lyu et al., 2020). Evidence has suggested that the relationship between Didemnum vexillum and Mytilus edulis reduces the predation by the green crab as Didemnum vexillum deters predator attacks (Auker et al., 2014). It was suggested that the negative impacts of Didemnum vexillum overgrowth on mussel growth, resulting in smaller-sized blue mussels, may protect smaller blue mussels from predation as these are preferred over larger blue mussels by predators (Auker et al., 2014). In Auker’s (2010) study, Carcinus maenas consumed fewer mussels that were overgrown by Didemnum vexillum. The toxic chemical defences of Didemnum vexillum and the release of secondary metabolites and sulfuric acid may deter crab predators (Lyu et al., 2020). The protection from predators provided by Didemnum vexillum may vary seasonally due to evidence suggesting the invasive ascidians deteriorate during the winter months, potentially reducing predation protection for mussels during this time (Auker et al., 2014).

However, Fletcher et al. (2013b) reported that smaller-sized Perna canaliculus mussels (20-40 mm) were significantly affected by fouling of Didemnum vexillum on cultured mussel ropes. The cultured ropes included ambient fouling (ropes left to be naturally colonized by Didemnum vexillum and other species), enhanced fouling (ambient fouling with ropes that were artificially inoculated with Didemnum vexillum) and a control (small levels of fouling maintained by the removal of Didemnum vexillum). The average mussel density and average mussel weight of smaller mussels were higher in the control than were in the treatments fouled by Didemnum vexillum. After 15 months, the smaller mussels were significantly smaller than the medium (40-60 mm) and large (60-70 mm)-sized mussels, which remained a similar size by the end of the experiment. They estimated there was a 40% reduction in small-sized mussel density per kilogram of Didemnum vexillum, indicating a negative relationship between small-sized mussel density and increasing Didemnum vexillum.

Small-sized mussels had a significant difference in mussel loss than the larger mussels, with greater loss of the smaller mussels seen in the ambient and enhanced fouling treatments. The small mussels were displaced and overgrown by Didemnum vexillum. Displacement was also evident to a lesser extent in the medium mussels but was less of a threat to larger mussels. However, the fouling treatments alone did not have a significant overall effect on mussel loss. It was suggested that high levels of fouling on the ropes may have resulted in small mussel loss as the mussels carry out a process of self-thinning, but high levels of fouling did not appear to affect individual mussel size or condition directly (Fletcher et al., 2013b). Fletcher et al. (2013b) also noted that Didemnum vexillum clogged cages and mesh used to house shellfish (e.g. mussels and oysters), which could reduce shellfish growth rates. Fletcher et al. (2013b) concluded that there were no direct effects of Didemnum vexillum fouling on mussel size and condition, but did indicate negative effects on small-sized mussels. However, in their study, Didemnum vexillum was only one of the fouling species contributing to fouling effects (Fletcher et al., 2013b).

Sensitivity assessment. In the Wadden Sea, Crepidula fornicata may out-compete and replace mussel beds in favourable winters.  Although no records of Crepidula fornicata replacing or dominating mussel beds in the UK were found, it has the potential to do so in both the subtidal and lower intertidal. Therefore, a precautionary resistance of ‘Low’ is suggested. Resilience is likely to be ‘Very Low’ as the slipper limpet population would need to be removed for recovery to occur. Therefore, sensitivity to invasion by Crepidula is assessed as ‘High’ but with 'Low' confidence due to the lack of direct evidence. 

The Pacific oyster Magallana gigas was reported to out-compete and replace mussel beds in intertidal sedimentary habitats. Magallana gigas may also develop reefs or grow on mussel beds in the upper subtidal but the evidence is less clear. Herbert et al. (2012, 2016) noted that blue mussels were found in areas dominated by Magallana gigas. But small clumps or occasional individuals would not constitute a blue mussel bed so the component biotopes would be lost. Similarly, mixed mussel and oyster beds have developed in the Wadden Sea over a period of 20 years (Reise et al., 2017). However, the resultant ‘oyssel’ beds have a lower density of mussels than pure mussel beds and represent a unique habitat so the component biotopes would be lost. Therefore, a precautionary resistance of ‘None’ is suggested for intertidal blue mussel beds and resilience is likely to be ‘Very low’ as the Magallana gigas population would need to be removed for recovery to occur. Therefore, sensitivity is assessed as ‘High’ for intertidal mussel bed biotopes on sediment.

The evidence presented shows Didemnum vexillum can overgrow mussels. In these biotopes, the mussels provide suitable hard substrata and stabilise the sediment for successful colonization of Didemnum vexillum, which may otherwise not colonize sandy and muddy sediments. Didemnum vexillum has been recorded in the lower intertidal but in the mid-shore examples of the biotope, the abundance and extent of colonies may be limited due to emersion. Didemnum vexillum colonies can survive exposure to air at low tides for a short time (not exceeding two hours) (Valentine et al., 2007a). The above evidence reported that mortality rates were low but that mortality probably varied between sites, the size of the mussels, and seasonally (Fletcher et al., 2013b). Therefore, a resistance of 'Medium' (some mortality, <25%) is suggested as a precaution. Resilience is likely to be 'Very low' as Didemnum vexillum would need to be physically removed for recovery to occur. Hence, sensitivity to invasion by Didemnum is assessed as 'Medium'. 

Introduction of microbial pathogens

Benchmark. The introduction of relevant microbial pathogens or metazoan disease vectors to an area where they are currently not present (e.g. Martelia refringens and Bonamia, Avian influenza virus, viral Haemorrhagic Septicaemia virus). Further detail

Evidence

Evidence for the impacts of microbial pathogens on Mytilus edulis was reviewed by Mainwaring et al. (2014) with specific reference to the shellfish pathogens Marteilosis and Bonamia. Natural Mytilus edulis beds are host to a diverse array of disease organisms, parasites and commensals from many animal and plant groups including bacteria, blue-green algae, green algae, protozoa, boring sponges, boring polychaetes, boring lichen, the intermediary life stages of several trematodes, copepods and decapods (Bower, 1992; Gray et al., 1999; Bower, 2010).

 Whilst Bonamia has been shown not to infect Mytilus edulis (Culloty et al., 1999), Marteilia refringens can infect and have significant impacts on the health of Mytilus edulis.  Its distribution, impacts on the host, diagnostic techniques and control measures are reviewed by Bower (2011). There is some debate as to whether there are two species of Marteilia, one which infects oysters (Marteilia refringens) and another that infects blue mussels (Marteilia maurini) (Le Roux et al., 2001) or whether they are just two strains of the same species (Lopez-Flores et al., 2004; Balseiro et al., 2007).  Both species are present in southern parts of the United Kingdom.  The infection of Marteilia results in Marteiliosis which disrupts the digestive glands of Mytilus edulis especially at times of spore release.  Heavy infection can result in a reduced uptake of food, reduced absorption efficiency, lower carbohydrate levels in the haemolymph and inhibited gonad development particularly after the spring spawning resulting in an overall reduced condition of the individual (Robledo et al., 1995).

Recent evidence suggests that Marteilia is transferred to and from Mytilus edulis via the copepod Paracartia grani.  This copepod is not currently prevalent in the UK waters, with only a few records in the English Channel and along the South coast.  However, it is thought to be transferred by ballast water and so localised introductions of this vector may be possible in areas of mussel seed transfer e.g. the Menai Strait.  The mussel populations here are considered to be naive (i.e. not previously exposed) and, therefore, could be heavily affected, although the likelihood is slim due to the dependence on the introduction of a vector that is carrying Marteilia and it then being transferred to the mussels.

Berthe et al. (2004) concluded that Mytilus edulis is rarely significantly affected by Marteilia sp.  However, occasions have been recorded of nearly 100% mortality when British spat have been transferred from a ‘disease free area’ to areas in France were Marteilia sp. are present.  This suggests that there is a severe potential risk if naive spat are moved around the UK from northern waters into southern waters where the disease is resident (enzootic) or if increased temperatures allow the spread of Marteilia sp. northwards towards the naive northern populations.  In addition, rising temperatures could allow increased densities of the Marteilia sp. resulting in heavier infections which can lead to mortality.

Sensitivity assessment. Bower (2010) noted that although Marteilia was a potentially lethal pathogen of mussels, most populations were not adversely affected by marteilioisis but that in some areas mortality can be significant in mariculture (Berthe et al., 2004).  The resultant population would be more sensitive to other pressures, even where the disease only resulted in a reduced condition.  Therefore, a precautionary resistance of ‘Medium’ is suggested (<25% mortality), with a resilience of ‘Medium’ (2-10 years) resulting in a sensitivity of ‘Medium’.  

Removal of target species

Benchmark. Removal of species targeted by fishery, shellfishery or harvesting at a commercial or recreational scale. Further detail

Evidence

Mytilus edulis is a commercially targeted species worldwide and has been fished for hundreds of years and managed in England and Wales for the last hundred years (Holt et al., 1998).  Mussels are collected on a commercial scale, in both the intertidal and subtidal, by dredges of various forms and by divers (Narvarte et al., 2011).  Damage caused by direct physical impacts which are assessed in under ‘abrasion’ and ‘penetration and/or disturbance of the surface of the seabed’ pressures, the sensitivity assessment for this pressure considers any biological effects resulting from the removal of target species on Mytilus edulis beds.

Mytilus edulis is the most important characterizing species defining the assessed biotopes and therefore any removal of the species will result in the removal of its associated fauna and a decline in species richness. Removal of most of the mussel biomass will also lead to loss of or reclassification of the biotope. The sensitivity to removal can be characterised as the immediate direct impact of harvesting and subsequent indirect effects.

Reports of dredging efficiency vary from 15% using Baird dredges on ground previously dredged for oysters (Palmer et al., 2007) to 90% using artisanal dredges (Narvarte et al., 2011). Mussels are also regularly hand collected by fisherman for bait and food from intertidal beds which can also result in significant damage to the bed (Holt et al., 1998; Smith & Murray, 2005).

Smith & Murray (2005) examined the effects of low-level disturbance and removal on an extensive bed of Mytilus californianus (composed of a single layer of mussels) in southern California.  They observed a significant decrease in mussel mass (g/m²), density (no./m²), percentage cover and mean shell length due to low-intensity simulated bait-removal treatments (2 mussels/month) for 12 months (Smith & Murray 2005).  They also stated that the initial effects of removal were ‘overshadowed’ by the loss of additional mussels during time periods between treatments, probably due to the indirect effect of the weakening of byssal threads attachments between the mussel leaving them more susceptible to wave action (Smith & Murray, 2005).  The low-intensity simulated bait-removal treatments had reduced percentage cover by 57.5% at the end of the 12 month experimental period.  Smith & Murray (2005) suggested that the losses occurred from collection and trampling are far greater than those that occur by natural causes.  This conclusion was reached due to significant results being displayed for human impact despite the experiment taking place during a time of high natural disturbance from El Niño–Southern Oscillation (ENSO). In addition, Holt et al., (1998) recorded an incident of the removal of an entire bed that is adjacent to a road in Anglesey due to fishermen bait collecting. 

Commercial removal of mussels can often be responsible for the depletion of mussel stocks.  For example, a substantial reduction in the mussel stock was observed in the Wash (England) during the 1990’s due to high fishing mortality and low recruitment (Atkinson et al., 2003).  The dredging fishery for mussels in the Limfjorden, Denmark, was reported to reduce the stock size of mussels (Dolmer et al., 1999).  The total stock of mussel in the Limfjorden was estimated to be 771 kt to 616 kt in 1993-1994, while the mean exploitation rate of the fishery was 14%.  In 1993-94 the size of mussel landings was found to correlate with a reduction in the overall stock size of the area, suggesting that mussel mortality was significantly increased by the fishery.  However, in 1995 the total stock had fallen to 494kt and the mean exploitation rate was 15% but there was no significant relationship between landings and stock size (Dolmer et al., 1999).  Divers observed that dense beds are likely to be more efficiently dredged due to their byssal attachments detaining the dredge bellow the carpet of the mussels whilst mussels in low-density beds cause the dredge to bounce along the seabed resulting in reduced efficiency (Dolmer et al., 1999).  A low level of exploitation may actually increase the growth rate of the mussels by reducing the intraspecific competition for food (Dolmer et al., 1999). However, Dolmer et al. (2001) observed that the mussel biomass was significantly lower in dredged areas suggesting that the lowering of the intraspecific competition does not increase the accumulation of biomass.

Recreational fishermen will often collect moulting Carcinus maenas or whelks by hand from intertidal mussel beds for bait.  The removal of predatory crabs could actively benefit the population this effect could be beneficial to mussel populations.

Sensitivity assessment Mytilus edulis beds have no avoidance mechanisms to escape targeted harvesting and as a result, a significant proportion of the bed can be removed (Palmer et al., 2007; Narvarte et al., 2011).  Dredging occurs on both subtidal and intertidal soft sediment and results in the removal of the mussel beds which defines the biotope.  As the majority of the mussel beds that are harvested in the UK are regularly replenished with seed, the recovery rate for maintained beds should be rapid.  In natural (wild) beds, the recovery could be significantly longer due to indirect effects from wave action and the sporadic nature of recruitment (Paine & Levin 1981; Seed & Suchanek 1992).  Mussel beds on hard substrata are unlikely to be affected by dredges and are therefore only vulnerable in the intertidal areas where they may be accessed for hand picking.  However, even hand-picking for bait can result in a significant decrease in cover, especially in beds composed of a single layer of mussels (Smith & Murray 2005).  It should be noted that dense, multi-layered mussel beds may be more resistant to the gaps and bait collection, as damage to the upper layer may not affect deeper layers, so that attachment to the substratum and each other is maintained (Brosnan & Crumrine, 1994). Based on the available evidence all Mytilus edulis bed biotopes are considered to have ‘Low’ resistance to this pressure and ‘Medium’ resilience so that sensitivity is assessed as ‘Medium’.

Removal of non-target species

Benchmark. Removal of features or incidental non-targeted catch (by-catch) through targeted fishery, shellfishery or harvesting at a commercial or recreational scale. Further detail

Evidence

Blue mussel beds may be removed or damaged by static or mobile gears that are targeting other species. The direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures, while this pressure considers the ecological or biological effects of by-catch. Removal of a large part of the Mytilus edulis bed unintentionally would significantly alter the biotope as the mussels are the key characterizing, structuring and functional species.

The removal of Mytilus edulis predators including the starfish Asterias rubens and Luidia ciliaris and the crabs Cancer pagurus and Necora puber as by-catch through commercial fishing activities could potentially benefit beds of mussels but the population of starfish and crabs are highly mobile and probably attracted to damaging and dying organisms left after dredging, and therefore likely to recover before the mussels are able to recruit (Gubbay & Knapman, 1999).  Fishing activities may exposure infauna, and leave dead and damaged species on the seabed, and areas where discards and by-catch have been deposited may also attract predators and scavengers.  But this potentially heightened level predation only lasts for a few days (Dolmer et al. 2001 observed a seven-day effect) and therefore is unlikely to have an impact on the bed as a whole.

Sensitivity assessment. The biogenic structure created by Mytilus edulis is the key characterizing, structural and functional feature of this biotope group. Removal of individuals as by-catch would substantially impact the biotope and hence this group is considered to have ‘Low’ resistance to this pressure and to have ‘Medium’ resilience. Sensitivity is, therefore, assessed as ‘Medium’.